an evaluation of the effects of wastewater treatment...
TRANSCRIPT
An evaluation of the effects of wastewater treatment
initiatives on water quality in coastal waters along the
Coral Coast, southwest Viti Levu, Fiji Islands
By
Exsley Jemuel TALOIBURI
A Thesis Submitted in Partial Fulfillment of the Requirements for
the Degree of Master of Science in Marine Science.
School of Islands and Oceans
Faculty of Science, Technology and Environment
The University of the South Pacific, Suva, Fiji
2009
You are encouraged to cite my thesis with proper citations and
acknowledgements.
Abstract
Most tourist hotels and villages on the Coral Coast of Fiji are situated
along the coastline resulting in higher observed coastal wastewater
pollution. This study was aimed at evaluating the effects of wastewater
treatment initiatives on water quality in coastal waters along the Coral
Coast, Fiji. Monitoring was undertaken on a planted gravel bed
constructed wetland at Tagaqe Village, Crusoe‘s Resort wastewater
treatment system, Coral Coast nearshore sites, Votua Village Creek, and
an ex-situ greywater treatment drum experiment. Results for the wetland
showed removal efficiency range of 94.7-99.3% for faecal coliform, E/coli,
total suspended solids (TSS) and biological oxygen demand (BOD).
Nitrogen elimination ranged between 50% for nitrite and 82.6% for
ammonia. Total Kjeldahl Nitrogen (TKN) declined by 75.5%, Total
Phosphorus by 69.1% and phosphate by 75.5%. The system at Crusoe‘s
Resort indicated removal range of 63.6-94.7% for faecal coliform, E/coli,
TSS and BOD. Nitrite was reduced by 50.9%; nitrate by 68.5%; ammonia
by 72.7%; and TKN by 50.1%. Total phosphorus was reduced by 60.5%
and phosphate by 70.7%. For the Coral Coast nearshore water quality,
results showed a mean salinity of 32ppt; temperature of 29.4ºC;
dissolved oxygen level of 6.10mg/L; conductivity 49.92mS/cm; nitrate
4.16µM; ammonia 2.09µM; nitrite 0.35µM and 0.43µM for phosphate
with a N:P ratio of 17. Votua Creek data showed that the lower housing,
bridge and creek mouth experience polluted wastewater discharges
relative to upper housing and the dam. The village and Mike‘s Diver tap
water were observed to be of unsafe drinking water quality standards,
without further treatment, but the housing tap water was safe. The
greywater treatment drum experiment showed general removal
efficiencies for all parameters in both mesocosms but varied considerably
between different loading regimes, sampling intervals and individual
water quality parameters.
Acknowledgements
The researcher is grateful to the Institute of Applied Science (IAS) for
funding this project as well as the researcher‘s postgraduate studies at
the University of the South Pacific. Sincere acknowledgement is extended
to Professor William Aalbersberg (principal supervisor and IAS Director);
Dr Milika Sobey-Naqasima (co-supervisor); Dr Chris Tanner and Dr Tom
Headley of the National Institute of Water and Atmospheric Research
(NIWA) in Hamilton, New Zealand; Mr Sarabjeet Singh, Dr Bale Tamata,
analytical and administrative staff of the IAS; and technicians from the
School of Marine Studies particularly Mr Jone Lima and Shiv Sharma as
well as other school support staff.
Appreciation is also offered to the chiefs and people in all villages
monitored along the Coral Coast; tourist hotel owners, managers and
staff along the Coral Coast; overseas visiting professionals and
volunteers; postgraduate student colleagues; immediate and extended
family members and friends; and other individuals or authorities that
contributed in one way or another to the successful completion of this
research project. Without assistance, this study would not have been
undertaken successfully.
Table of Contents Page
Abbreviations i
List of Tables ii-iii
List of Figures iv-vii
Chapter 1 Introduction 1-17
1.1. General 1-5
1.2. Importance of wastewater management 6-7
1.3. Contents of wastewater that cause problems 8-11
1.4. Status of wastewater legislation in Fiji 11-12
1.5. Wastewater treatment practices 12-14
1.6. Implications of the study 14-15
1.7. Research objectives 15-16
1.8. Organisation of thesis 16-17
Chapter 2 Background of Fiji & the Coral Coast 18-39
2.1. Introduction 18
2.2. Fiji Islands and Viti Levu 18-25
2.3. The Coral Coast 26-30
2.4. Water quality standards for normal coral growth 31-33
2.5. Status of water quality in Fiji & the Coral Coast 33-39
Chapter 3 Nutrient enhancement & coastal waters 40-55
3.1. Introduction 40
3.2. Importance of coastal aquatic systems 40-42
3.3. Potential effects of nutrient enrichment on corals 43-55
Chapter 4 Monitored wastewater treatment systems 56-91
4.1. Introduction 56
4.2. Constructed wetlands for wastewater treatment 56-72
4.3. Commercial wastewater treatment systems 72-85
4.4. Greywater treatment drum experiment 85-91
Chapter 5 Methodology 92-108
5.1. Introduction 92
5.2. Field sampling procedures 92-95
5.3. Greywater treatment drum experiment 96-97
5.4. Analytical methods 97-100
5.5. Automated Flow Injection Analysis 100-108
Chapter 6 Results 109-128
6.1. Introduction 109
6.2. Tagaqe Village constructed wetland 109-111
6.3. Crusoe’s Resort treatment plant 111-112
6.4. Coral Coast water quality monitoring 112-117
6.5. Votua Creek water quality monitoring 118-121
6.6. Drum system model experiment 122-128
Chapter 7 Discussion 129-158
7.1. Tagaqe Village constructed wetland 129-135
7.2. Crusoe’s wastewater treatment system 135-140
7.3. Coral Coast nearshore water quality 140-147
7.4. Votua Creek water quality 147-151
7.5. Drum system model experiment 152-158
Chapter 8 Conclusions 159-164
Bibliography 165-183
Appendices 184-195
Appendix A: Tagaqe wetland monitoring data 184-186
Appendix B: Crusoe’s system monitoring data 187-188
Appendix C: Coral Coast monitoring data 189-193
Appendix D: Votua Creek monitoring data 194-195
i
Abbreviations
kW kiloWatts BOD Biological Oxygen Demand
TSS Total Suspended Solids DO Dissolved Oxygen TDS Total Dissolved Solids
COD Chemical Oxygen Demand FJ$ Fijian dollars
rtPBRs Recirculation textile Packed Bed Reactors mg/L milligram per litre ENSO El Nino Southern Oscillation
RSF Recirculation Sand Filter SPCZ South Pacific Convergence Zone EIA Environment Impact Assessment
WHO World Health Organisation STP Sewage Treatment Plant
UNEP United Nations Environment Programme mm/day millimetre per day m3 cubic metres
ppt parts per thousand m metre µg/L microgram per litre
µmol/L micromole per litre mmol/L millimole per litre
c/100ml counts per 100 millilitres NH3-N Ammonia Nitrogen NO3-N Nitrate Nitrogen
NO2-N Nitrite Nitrogen PO4-P Phosphate Phosphorus
TP Total Phosphorus TIN Total Inorganic Nitrogen TKN Total Kjeldahl Nitrogen
ºC degree Celsius ± plus or minus > greater than
< less than % percent
cm centimetre HLR Hydraulic Loading Rate APHA American Public Health Association
L Litres FIA Flow Injection Analyser
ml milliliter m2 square metre
ii
List of Tables
Table 1: page 29
List of general Coral Coast water quality sampling sites
Table 2: page 33
Summary of recommended standards for nearshore waters to support
coral reefs and recreation in Australia and New Zealand
Table 3: page 39
Comparison of the mean faecal coliform counts around Suva nearshore
waters and rivers from different studies
Table 4: page 39
Water quality results for the Port of Suva in 1992 as observed by Tamata
et al.
Table 5: page 91
Summary of typical greywater characteristics targeted in the experiment
Table 6: page 96
Preliminary water quality monitoring program to compare different
loading regimes
Table 7: page 97
Starting recipe for artificial greywater
Table 8: page 98
Accuracy and precision for each Lachat Quick Chem FIA method
Table 9: page 99
Method detection limit for each Lachat Quick Chem FIA method
Table 10: page 110
Summary of water quality results from Tagaqe wetland over the period
between June 2005 and October 2006
Table 11: page 111
Summary of water quality results from the Crusoe‘s wastewater
treatment plant over the period between October 2005 and September
2006
iii
Table 12: page 114
Summary of water quality results from the Coral Coast between July
2005 and July 2006
Table 13: page 117
Summarised ―baseline‖ water quality data from the Coral Coast over a
five year period prior to July 2005
Table 14: page 121
A summary of Votua Creek water quality monitoring between June and
September 2006
Table 15: page 123
Summary of results for ―Monitoring Period 1 – Large Doses‖
Table 16: page 124
A summary of results for ―Monitoring Period 2 – Moderate Doses‖
Table 17: page 126
Mean results for ―Monitoring Period 3 – Small Doses‖
iv
List of Figures
Figure 1: page 1
A village and associated latrine along the Coral Coast of Fiji, within ~10
m of beach high tide level
Figure 2: page 2
A piggery about 2m from high tide mark at Votua Village
Figure 3: page 2
Black spots of algal cover in front of Namaqumaqua Village along the
Coral Coast of Fiji
Figure 4: page 3
Algal proliferation along the Coral Coast in Fiji
Figure 5: page 3
Seaweed and seagrass breakage physically pollutes the beach in front of
the Fijian Shangri La Resort along the Coral Coast of Fiji
Figure 6: page 4
Exposed shoreline in front of Votua Village on the Coral Coast of Fiji due
to coastal erosion
Figure 7: page 5
(a) Wetland at Tagaqe; (b) wastewater treatment system at Crusoe‘s
Retreat; and dry composting toilets at Komave (c) and Tagaqe (d)
Figure 8: page 7
Discharge of untreated black water into a river at Votua Village
Figure 9: page 13
A household septic tank exemplifying ―primary treatment‖
Figure 10: page 14
A basic flow chart of the wastewater treatment train
Figure 11: page 19
Location of the Fiji Island group
v
Figure 12: page 28
Map of Coral Coast showing the location of the two primary wastewater
treatment systems being monitored
Figure 13: page 29
Location of general Coral Coast sampling sites, villages and hotels
Figure 14: page 41
Mangroves are a classic example of a coastal aquatic system
Figure 15: page 58
A systematic diagram of a horizontal flow constructed wetland
Figure 16: page 59
Diagram of a ―surface flow‖ constructed wetland
Figure 17: page 59
Diagram of a ―subsurface flow‖ wetland
Figure 18: page 63
A simplified diagram of the nitrogen processes and the flows of different
nitrogen forms in a wetland
Figure 19: page 66
Factors affecting the biological processes of denitrification on different
spatial scales
Figure 20: page 71
Construction stages of the gravel bed wetland at Tagaqe Village
Figure 21: page 72
Similar cross section diagram of Tagaqe Village wetland
Figure 22: page 79
The AdvanTex AX100 Treatment System at Crusoe‘s Retreat
Figure 23: page 80
Biotube Effluent filters used in septic tanks at Tagaqe and Crusoe‘s
Retreat
Figure 24: page 81
Schematic as built of the Crusoe wastewater treatment system
vi
Figure 25: page 84
(a) The ProSTEP Effluent pump switchboard at the recirculation tank; (b)
the lower septic pumping system closer to the beach at Crusoe‘s Resort
Figure 26: page 84
(a) George Reece standing beside the Carbon filter and ventilation fan of
the treatment system at Crusoe‘s; (b) The AX100 textile filter pod fibre
layers for wastewater treatment at Crusoe‘s Retreat
Figure 27: page 85
(a) The recirculation splitter valve and the effluent pumping system; (b)
the recirculation splitter valve with piping connections from septic tank
effluents and the AX100 treatment system
Figure 28: page 85
The flower gardens and ground disposal area at Crusoe‘s Retreat
Figure 29: page 88
(a) An in situ greywater treatment drum system at Votua Village along the
Coral Coast; and (b) an ex-situ model at the university
Figure 30: page 89
Side view of greywater treatment mesocosm
Figure 31: page 90
Details of drainage holes in 20L buckets used to contain coconut shell
and husk
Figure 32: page 94
Acid bath for field sampling bottles and reagent/standard preparation
Figure 33: page 94
Field sampling at Crusoe‘s wastewater treatment system
Figure 34: page 94
(a) Sample collection at the Tagaqe wetland inlet; (b) sample collection at
the Tagaqe wetland outlet
vii
Figure 35: page 100
The Auto Sampler Injector which sucks sample to be passed through the
FIA manifold
Figure 36: page 102
(a) The injector and sample zone; (b) reagents being added to samples; (c)
analyser pumps; (d) a 4 channel manifold; (e) nitrate column on
manifold; (f) the computer system that log results; (g) FIA waste outlet;
(h) peak shaped signals on computer for analyte
Figure 37: page 103
Schematic diagram of a typical flow injection analysis manifold
Figure 38: page 110
Sample of treated and untreated wastewater from Tagaqe wetland
Figure 39: page 122
Sampling the ex-situ drum system experiment
Figure 40: page 127
A graph showing the effluent flow rate vs. time for the two mesocosms
Figure 41: page 128
Some degree of clogging on the coconut husk layer within the High
Loading mesocosm
1
Chapter 1 Introduction
1.1. General
Most tourist hotels and resorts, villages, and industrial developments in
Fiji are situated along coastlines (Figures 1 & 2) to enable tourists,
villagers and industries to utilise the coral reef and lagoon environments
for activities such as swimming, transportation, fishing, snorkelling,
scuba diving and food collection (Thaman and Sykes, 2005).
Figure 1: A village and associated latrine (circled) along the Coral Coast of Fiji,
within ~10 m of beach high tide level [Tanner and Gold, 2004]
Recently high levels of seaweed growth (Figure 3) have been noticed
around some resorts and coastal areas throughout Fiji, particularly in
the Mamanuca Islands and along the Coral Coast, which is an indication
of elevated nutrient levels in the coastal water (Lovell and Tamata, 1996;
Vuki et al., 2000; Thaman and Sykes, 2005).
2
Studies conducted on water quality indicated high nutrient levels in
coastal waters (Mosley and Aalbersberg, 2003), above guidelines for coral
reef areas, and in some places high faecal coliform levels, an indication of
sewage pollution (Hodgson, 1999; Coral Cay, 2001; Coral Cay, 2005).
Both these phenomena are likely to lead to the degradation of coral reefs
and deteriorating water quality, which will be detrimental in the long
term to tourism and health of people in these respective areas.
Figure 2: A piggery about 2m from high tide mark at Votua Village
Figure 3: Black spots of algal cover in front of Namaqumaqua Village along the
Coral Coast of Fiji
3
The obvious Sargassum algal overgrowth in lagoons along the Coral
Coast of Fiji is a major concern given the importance of the region for
local communities and as a tourist destination. A large number of
tourists come to Fiji to enjoy tropical reefs, fish biodiversity, and scuba in
clear and clean water (Figure 4). Hence if the Coral Coast reefs and
nearshore ecosystems are degraded, it will result in adverse impacts for
hotel owners and local villagers that rely on tourism for income and
employment. Breakage of seaweed and seagrass can pollute pristine
beaches that resorts and hotels rely on for tourism attraction (Figure 5).
Figure 4: Algal proliferation along the Coral Coast in Fiji (Tanner & Gold, 2004)
Figure 5: Seaweed and seagrass breakage physically pollutes the beach in front of
the Fijian Shangri La Resort along the Coral Coast of Fiji
4
In addition fish stocks that many coastal dwellers harvest from
nearshore areas for protein will also be limited. Moreover coastal erosion
along the Coral Coast (Figure 6) is likely to increase as the reefs are
broken down by waves and not regenerated (Mosley and Aalbersberg,
2003).
Figure 6: Exposed shoreline in front of Votua Village on the Coral Coast of Fiji due
to coastal erosion
Immediate strategies are needed at community, regional and government
levels to try and reduce wastewater discharges and nutrient enrichment
on the Coral Coast of Fiji. An integrated approach to coastal management
is needed to manage and control land based sources of wastes before
being disposed into the coastal environment.
As a result sanitation engineers (Tanner and Gold, 2004) in collaboration
with the University of the South Pacific and other stakeholders studied
the nutrient pollution along the Coral Coast as part of the Fiji integrated
coastal management project and recommended:
5
a) Upgraded hotel sewage treatment by on-site systems;
b) Better village treatment using septic tank filters, dry composting
toilets, and/or artificial constructed wetlands;
c) In-situ composting of pig waste.
Since then, several model projects along these lines had been initiated
which include a sub-surface gravel bed wetland at Tagaqe Village; an
AdvanTex wastewater management system at Crusoe‘s Retreat;
composting dry toilets in Vunisinu Village in Rewa, Tagaqe and Komave
Villages in Nadroga; and a composting pig waste in one piggery at the
National Youth Training Centre in Sigatoka (Figure 7).
Figure 7 (a): Wetland at Tagaqe; (b) wastewater treatment system at Crusoe’s
Retreat; and dry composting toilets at Komave (c) and Tagaqe (d)
(a) Wetland at Tagaqe Village (b) Treatment system at Crusoe’s
(c) Dry composting toilet at Komave (d) Composting toilet at Tagaqe
6
1.2. Importance of wastewater management
Wastewater is liquid or water associated with waste disposed from
residences, institutions, commercial and industrial establishments,
together with groundwater, surface and storm water. Wastewater
includes dissolved contaminants, suspended solids and micro-organisms
(New Zealand Ministry for the Environment, 2003). Every community
produces both solid and liquid waste and the liquid portion is essentially
the water supply after it has been contaminated by the various uses to
which it has been exposed. Wastewater can be classified into four
categories including domestic (wastewater from residences and
commercial facilities); industrial (wastewater from industrial waste);
Infiltration (wastewater from leaking joints, cracks, porous walls, storm
drain connections, roof headers, manhole covers); and storm water
(wastewater runoff from flooding due to rainfall).
Wastewater from tourist resorts, villages and agricultural runoff
comprises greywater and black water. Greywater includes wastewater
from laundry, kitchen, shower and sink water. Black water consists of
sewage effluent, which can be the principal source of pollution in
nearshore waters if inadequately treated (Thaman and Sykes, 2005).
Likely problems associated with untreated wastewater effluent being
discharged into the coastal environment may include the following:
7
a) High Biological Oxygen Demand (BOD) – a measure of the amount
of oxygen that would be taken up by degrading the organic matter
in the effluent. Excessive BOD may result in low oxygenated waters
and fish kills.
b) High levels of nutrients, particularly nitrogen and phosphorus –
promotes excessive growth of plants, including algae and seaweed,
which can smother corals.
c) Pathogens (faecal coliform) – disease causing microorganisms that
can be harmful to swimmers and other organisms.
d) Suspended solids – particles that lead to poor water clarity and
smothering of corals.
The discharge of untreated greywater and sewage effluent (Figure 8) into
the environment can have major impacts on coastal habitats and human
health. Sources of pollution into coastal waters include discharge outlets
for greywater, seepage from septic tanks and soak pits, seepage from pit
or drum toilet systems, seepage from gardens and golf courses,
agricultural and fertiliser runoff through rivers, and sewage effluent
discharge outlets (Thaman and Sykes, 2005).
Figure 8: Discharge of untreated black water into a river at Votua Village
8
1.3. Contents of wastewater that cause problems
The impact of wastewater discharges on the environment depends on the
standard of wastewater discharged, volume discharged and
characteristics of the receiving environment.
1.3.1. Organic material
Wastewater organic contents consist of human faeces, protein, fat,
vegetable and sugar material from food preparation, and washing
detergents. Some of this is dissolved into the water whilst some exists as
separate particles.
In a natural system, bacteria from the soil and water usually consume
organic materials from wastewater to promote growth. In a healthy
diluted water environment where there is adequate dissolved oxygen,
aerobic (oxygen-using) bacteria tend to feed on the organic material and
form a slime of new bacterial cells and dissolved salt waste products. On
the contrary, if undiluted wastewater is isolated anaerobic (non-oxygen-
using) bacteria would decompose the waste organic material in the
process releasing odorous gases such as hydrogen sulphide and other
non-smelly gases such as methane and carbon dioxide. Thus, it is the
amount of oxygen removed or the rapid growth of bacterial slime that can
cause harmful effects on the coastal environment (New Zealand Ministry
of Environment, 2003).
9
In addition where there is an overwhelming amount of wastewater,
available dissolved oxygen will be used up resulting in deoxygenated
water columns. This can have adverse impacts on fish and other forms of
oxygen-dependent life. As a result, it is equally important for wastewater
to be treated in order to reduce as much organic material as possible.
1.3.2. Suspended solids
The portion of organic material that does not dissolve but remains
suspended in water is known as suspended solids. When untreated
effluent is discharged into rivers or a water body, accompanying solids
will tend to settle in quiet spots where there is no or little water flow. In
extreme cases, this would result in an anaerobic condition, which can be
harmful to fish, and other oxygen dependent life forms at the bottom of
streams and creeks.
1.3.3. Dissolved salts
The most significant salts in wastewater are nitrates and phosphates.
These naturally occur in coastal waters to some extent. Nitrate is derived
from the breakdown of organic nitrogen in protein waste matter, and the
oxidation of ammonia in urine. Phosphates are present in detergents
used in washing and laundry, and are also produced by organic
breakdown. Nitrates and phosphates are essential elements for growth.
When nitrates and phosphates are discharged into natural waters they
fertilise the growth of microscopic algae and seaweeds.
10
Coral reefs flourish in clean, nutrient poor waters and are very sensitive
to changes in their environmental conditions such as increased levels of
freshwater, sediment and nutrients (Castro and Huber, 2003). Slight
increases in nitrate levels in coastal waters can lead to growth of dense
mats of algae and seaweed on reef areas. The algae often overgrow and
smother the reef, preventing fish and other reef inhabitants from finding
food and shelter (Goreau and Thacker, 1994; McCook, 1999). Increases
in phosphate levels can lead to brittleness of coral and crumbling (Mosley
and Aalbersberg, 2003).
1.3.4. Bacteria and viruses
The human gut produces a huge quantity of bacteria, which are excreted
as part of faeces on a daily basis. The most common and easily measured
organism is the faecal coliform bacterium. This is called an indicator
because its presence indicates the presence of faecal matter from warm-
blooded animals.
The discharge of non-disinfected sewage effluent into the marine
environment may result in bacterial contamination of waters and
organisms. Many of the faecal coliform bacteria in human waste are
harmless. However, there are disease organisms or pathogens than can
cause harm to human health. These include bacteria such as typhoid, or
viruses such as hepatitis B. Direct contact with these pathogens or
11
pollution of the water supply can result in infections. This poses a public
health risk (i.e. nausea, vomiting, diarrhoea, ear, and throat infections)
to people who use the waters for recreation or fish harvesting (New
Zealand Ministry of Environment, 2003).
Moreover relatively high concentrations of pathogens in water can also
make an area unsafe for swimming and recreation. A likely effect of
direct contact with polluted water is skin irritation and scratchiness.
Therefore it is important that levels of faecal coliform do not exceed
recreational exposure standards.
1.4. Status of wastewater legislation and regulation in Fiji
Current legislation or regulations on waste discharges into the
environment in Fiji are either inadequate or absent. Subsequently there
is no agency that consistently monitors the quality of Fiji‘s coastal
waters. However there are some upgraded existing waste disposal
practices which ensure that new tourism developments undergo
Environmental Impact Assessments (EIA) that encourage proper waste
treatment.
The new Environment Management Bill (2004) which was enacted by the
Parliament of Fiji on 17 March 2005 requires a tourism facility to obtain
a permit to discharge waste or pollutants into the environment (Article
35) and fines up to FJ$250,000 if the facility is found to be polluting
12
(Article 45). The Bill also requires a tourism facility to establish an
environmental management committee (Article 16). Thus, it is in the best
interest of prospective stakeholders to invest in upgrading their
wastewater treatment systems and undertake environmental monitoring
not only in terms of improving their surrounding environment but also to
comply with the new legislation.
1.5. Wastewater treatment practices
Wastewater management practices have different stages including
primary, secondary and tertiary treatments. The initial treatment of
wastewater is referred to as ‗primary treatment‘, which focuses on the
separation of solids from liquid in order to reduce suspended solids and
biological oxygen demand from wastewater effluents. Good examples of
this stage include settling tanks in sewage treatment plants and
household septic tanks (Figure 9). The process allows solids to settle at
the bottom of the septic tank while organic matter is digested by
bacteria. Sludge removal depends on the size and level of the tank. From
this stage, pathogens are filtered and ammonia-like material is converted
to nitrate through nitrification bacteria (Castro and Huber, 2003).
However since nitrate can flow easily through sand and groundwater
without degrading, only 15 percent of nitrogen from wastewater is
removed from septic tanks. Thus primary treatment alone is insufficient
(UNEP, 2002).
13
Figure 9: A household septic tank exemplifying “primary treatment”
The second stage of effluent treatment can be defined as ‗secondary
treatment‘. This process aims to remove biological oxygen demand (BOD)
and bacteria. Widely used examples that utilise aerobic bacteria to
decompose organic matter include trickling filters (wastewater is sprayed
over rocks or plastic media that are coated with a slimy layer of bacteria),
activated sludge (wastewater mixed with bacteria containing sludge and
air), and oxidation/algal ponds (large ponds where organic matter is
consumed by bacteria using oxygen which is supplied algal growth)
(Castro and Huber, 2003; Thaman and Sykes, 2005). Another classic
example is constructed wetlands, which provide secondary treatment
through advanced nutrient removal as a result of denitrification (UNEP,
2002).
14
Tertiary treatment is the third stage of wastewater management. Tertiary
treatment covers a whole range of processes including disinfection,
nutrient removal and removal of anything that has not been dealt with at
the secondary stage but is deemed necessary to get out. Classic examples
are tertiary ponds and wetlands (UNEP, 2002). In certain cases, before
treated effluent is discharged into the coastal environment or sprayed
onto land vegetation the effluent is further disinfected using ultraviolet
light, ozone, or chlorine (Thaman and Sykes, 2005).
Figure 10: A basic flow chart of the wastewater treatment train
1.6. Implications of the study
This study was derived purposely to monitor the efficiency of specific
village and hotel wastewater treatment initiatives along the Coral Coast
of Fiji, particularly the subsurface flow constructed wetland at Tagaqe
Village and the wastewater management system at Crusoe‘s Retreat to
determine their likely effects on nutrients in coastal water and viability
for future development in other villages and tourist resorts throughout
Fiji and the Pacific. Results from this study will act as baseline
information on the performance of wastewater treatment systems
15
relevant to coastal villages and tourist resorts in Fiji and elsewhere
throughout the Pacific.
Cost-effective treatment plants, such as the constructed subsurface flow
wetland at Tagaqe Village and the wastewater treatment plant at
Crusoe‘s Retreat are the first of their kinds to be constructed and
monitored in Fiji. If they are effective and viable then there are
possibilities that they can be promoted and further developed in other
Pacific Island countries.
1.7. Research objectives
The central aim of this project was to evaluate the effects of wastewater
treatment initiatives on water quality in coastal waters along the Coral
Coast of Fiji. This was accomplished with the following objectives:
a) To determine influent and effluent water quality parameters including
temperature, salinity, dissolved oxygen, total suspended solids, pH,
conductivity, biological oxygen demand, and coliform levels from
wastewater treatment systems.
b) To evaluate nutrient levels from untreated and treated outlets
especially nitrate, nitrite, ammonia, total nitrogen, total Kjeldahl
nitrogen, phosphate and total phosphorus from the Tagaqe Village
constructed wetland and Crusoe‘s Retreat wastewater treatment plant
wastewater treatment systems to determine the levels of removal
efficiency.
16
c) To compare water quality trends and nutrient levels in Coral Coast
nearshore waters adjacent to tourist hotels and villages. This was
achieved by comparing data for this present study with baseline
monitoring data obtained by the Institute of Applied Science
researchers.
d) To analyse the actual pollution level along the Votua Village Creek, as
a typical Coral Coast freshwater stream that is utilised for piggery
farming, with relatively considerable wastewater discharge from both
human and animal sources.
e) To set up an ex-situ greywater treatment model drum experiment
mimicking on-site drum trial at Votua Village with different substrates
including gravel, sandy soil, coconut husk and shells, and coral
rubble to be dosed with known concentrations of prepared artificial
greywater solutions, and then analyse effluent samples from the drum
systems, in order to understand the potential efficiency in an in-situ
system.
1.8. Organisation of the thesis
This section outlines the thesis overview in order to assist readers to
locate appropriate chapters, which are of interest to them. The thesis
consists of eight major chapters excluding other sections such as the
Abstract, Acknowledgement, Dedication, List of Figures, List of Tables,
Abbreviations, Table of Contents, Bibliography and Appendices. Chapter
1 entails the Introduction; Chapter 2 consists of a Background of Fiji
17
Islands and the Coral Coast in relation to water quality status; Chapter 3
comprises the Literature review of nutrient enrichment impacts on
aquatic systems and water quality standards for normal coral growth;
Chapter 4 describes the different wastewater treatment initiatives
monitored; Chapter 5 contains the Methodology; Results in Chapter 6;
Discussion section in Chapter 7; and Conclusions in Chapter 8.
18
Chapter 2 Background of Fiji and the Coral Coast
2.1. Introduction
This chapter reviews the background of the Fiji group, particularly the
study site on the Coral Coast of Fiji. Aspects covered include location,
coastal zone, climate and hydrology, human impacts on coastal
environments, water quality standards for normal coral growth and
recreation, and the status of water quality in Fiji.
2.2. Fiji Islands and Viti Levu
2.2.1. Location
Fiji is an island nation in the South Pacific Ocean which occupies an
archipelago of more than 322 islands excluding atolls and reefs, of which
106 are permanently inhabited, and 522 smaller islets (Vuki et al., 2000;
Singh, 2001). The Fiji group is situated between 15° and 22°S latitude
and 174°E and 177°W longitude (Figure 11).
Most of Fiji‘s land area consists of two large mountainous islands that
account for 87 percent of the country‘s total human population. The two
major islands are Viti Levu with an area of 10,400 km2, and Vanua Levu
with an area of 5,540 km2 (Watling and Chape, 1992). The capital of Fiji
Islands is Suva, which is located on the southeastern side of Viti Levu.
19
Figure 11: Location of the Fiji Island group
2.2.2. The coastal zone of Fiji
The Fiji Islands is surrounded by a vast maze of reefs, particularly on the
southwest, northwest and northeast coasts of Viti Levu and extending to
other locations throughout the Fiji group. Well developed barrier reefs
are common around the many islands off the northwest coast of Viti
Levu. Sediments produced are largely derived from broken corals,
calcareous algae, molluscan fragments and foraminifera (Maharaj, 1998;
Singh, 2001).
Similarly, the shallow coastal zone of Fiji comprises of three major,
interrelated habitat types: marine algae and seagrass; large areas of
20
mangroves; and extensive coral reefs. The marine resources include
approximately 1000 coral reefs with representatives of all major reef
types (Vuki et al., 2000). Although marine biodiversity is lower than in
the ‗coral triangle‘ of Indonesia, the Philippines, Papua New Guinea and
northeastern Australia, Fiji does support approximately 200 species of
coral (Veron, 2000). Furthermore it has been estimated that Fiji has
approximately 1200 marine fish species (Vuki et al., 2000).
Fiji‘s population in 2000 was above 775,000 and increasing rapidly
(South and Skelton, 2000). Since much of this population is
concentrated around the coast, the expanding development of coastal
areas and exploitation of the reefs are resulting in a suite of threats to
the coral reefs including siltation, eutrophication and pollution (Coral
Cay, 2005). For example, some of the natural landscape has been
converted for agriculture, particularly sugar cane, which impacts the
coastal environment via soil erosion leading to elevated sediment loads
smothering coral colonies. Further erosion is also caused by the removal
of mangroves to reclaim land for urban development. Such expansion of
urban areas has also led to pollution of the coastal zone because of
inadequate sewage treatment and waste disposal. Industrial point
sources have also been shown to contribute to decreasing water quality
(Coral Cay, 2001).
21
Specifically Viti Levu has about 750 km of coastline, of which over 94
percent comprises fringing reefs and mangroves. The remaining 6
percent consists of open or exposed coast, without natural reef or
mangrove protection (Maharaj, 1998). Moreover, half of Viti Levu‘s
coastline entails undeveloped areas with mangroves and reef fringes with
low near sea elevations of less than 3 m above mean sea level. Coastal
settlements and agricultural developments account for about 23 percent
of the Viti Levu coastline (Maharaj, 1998).
2.2.3. Climate and hydrology
The countries within the Pacific Ocean experience a variety of weather
and climate due to their wide ranging geographical locations, which
comprise both tropical and semi-temperate latitudes (Rapaport, 1999).
The dominant climate feature that affects the Fiji region is the South
Pacific Convergence Zone (SPCZ), a zone associated with high rainfall,
which fluctuates northeast and southwest (Salinger et al., 1995).
Fiji has a tropical maritime climate that is generally pleasant year-round,
lacking excessive temperature variation. Due to the easterly and
southeasterly trade winds predominance, the climatic conditions vary
from moderately hot and moderately dry on the leeward side of Viti Levu
to warm and wet on the windward side of the island (Pahalad, 1995).
22
During all seasons the common winds over Fiji are the trade winds from
the east to southeast. On the coast of the two main islands, Viti Levu and
Vanua Levu, day-time sea breezes blow across with great regularity.
Winds over Fiji are generally light or moderate; stronger winds are far
less common and are most likely to occur in the period June to
November when the trade winds are most persistent. However, tropical
cyclones and depressions can cause high winds, especially from
November to April when the trades die down (Koushy and Leetmaa,
1989).
In Fiji there are two major seasons including the wet, hot season lasting
from November through April and the warm, dry season, lasting from
May through October. The main island of Viti Levu is drier and more
temperate on the western side (e.g. Sigatoka to Rakiraki). The eastern
part where Suva is located is known for its wet and cloudy weather. The
islands off the coast are generally sunny and more stable than the
mainland particularly the Mamanuca Group off the west side of Viti
Levu.
At lower levels around Fiji the air temperatures are fairly uniform. In the
lee of the mountains on the largest islands however, the day time
temperatures are often 1° to 2°C above those on the windward sides.
Also, the humidity on the lee side tends to be somewhat lower. Due to
23
the influence of the surrounding ocean, the changes in the temperature
from day to day and season to season are relatively small. The average
temperatures change only about 2° to 4°C between the coolest months
(July and August) and the warmest months (January to February).
Around the coast, the average night time temperatures can be as low as
18°C and the average day time temperatures can be as high as 32°C. In
the central parts of the main islands, average night-time temperatures
can be as low as 15ºC (Fiji Meterological Service, 2004).
Similarly, water temperature fluctuations in Fiji are negligible averaging
around 26oC throughout the year making its crystal blue waters perfect
for snorkelling and other water activities.
With reference to rainfall, it is highly variable and strongly influenced by
topography, with the prevailing southeast trade winds bringing moisture
onshore and causing heavy showers in the mountain regions. These
trades are often saturated with moisture, and any high landmass lying in
their paths receive much of the precipitation. The mountains of Viti Levu
and Vanua Levu create wet climatic zones on their windward sides and
dry climatic zones on their leeward sides resulting in wet and dry zones
with that are fairly well defined. On the outer islands and other small
islands nearby the climatic differences from one part to another of
individual islands is insignificant (Fiji Meteorological Service, 2004).
24
Fiji‘s wet season is controlled largely by the north and south movements
of the South Pacific Convergence Zone, the main rainfall producing
system for the region. The wet season is characterised by heavy, brief
local showers and contributes most of Fiji's annual rainfall. Rainfall is
usually abundant during the wet season, especially over the larger
islands, and it is often deficient during the rest of the year, particularly
in the "dry zone" on the northwestern sides of the main islands (Fiji
Meteorological Service, 2004).
Annual rainfall in the dry zones averages around 2000mm (79 inch),
whereas in the wet zones, it ranges from 3000mm (118 inches) around
the coast to 6000mm (236 inches) on the mountainous sites. The smaller
islands receive various amounts according to their location and size,
ranging from around 1500mm (59 inches) to 3500mm (138 inches). The
southeastern parts of the main islands, generally receive monthly total
rainfall of 150mm (6 inches) during the dry season, and 400mm (16
inches) during the wettest months. These parts of the islands have rain
on about six out of ten days for the dry season, and about eight out of
ten days for the wet season. The northwestern parts of these islands are
in the rain shadow and receive generally less than 100mm (4 inches) per
month during the dry period. The variation in the monthly totals between
the two zones during the wet season is little (Fiji Meteorological Service,
2003 & 2004).
25
On average, around 10 to 15 cyclones per decade affect some part of Fiji
with a few causing severe damage. Specific locations may not be directly
affected for several years but the dominant northwest to southeast
cyclone track gives some increased risk of damage in the outlying
northwest island groups. Large-scale flooding in Fiji is mostly associated
with the passage of a tropical cyclone or depression resulting in
prolonged heavy rainfall. Normally urban centres situated near the
mouth of the four main rivers on the main island are affected the most.
Localised flash flooding during the wet season is common on a small
scale. Storm tides and heavy swells can also result in flooding of low-
lying coastal areas during the pass of a severe cyclone (Fiji Meteorological
Service, 2004).
Moreover, droughts in Fiji can be closely linked to the ENSO (El Nino
Southern Oscillation) phenomenon, which results in generally below
average rainfall for Fiji. A strong ENSO episode is likely to result in a
major drought over the country, as happened during 1982 - 1983 and
1997 - 1998 ENSO events. Otherwise, even in a normal year the rainfall
in the "dry zones" of the country is so low during the Dry Season that an
incident of below average rainfall for a few months can cause a drought
effect (Koushy and Leetmaa, 1989).
26
2.3. The Coral Coast
2.3.1. Location of study sites
The principal study site for this research was along the Coral Coast. The
Coral Coast is the name given to the southwest coastline of Fiji‘s largest
island, Viti Levu Island, which is a popular tourist retreat and
destination for a Fiji vacation. It is located about 190 kilometres from
Suva on the Queen's Highway between Nadi and Suva (Coral Cay, 2005).
Along the Coral Coast are a handful of large resorts (including the Fijian
Shangri-La, Naviti, Warwick, Hideaway, Crusoe‘s and Outrigger) and
many medium and small resorts. Between the resorts are large stretches
of uninhabited rainforest with occasional coastal villages. The coastline is
a combination of bays, reefs, beaches, rocky outcrops and mangrove
forests. Sigatoka is the Coral Coast's main town.
For the purpose of this study, the Coral Coast is only defined as the
coastline between Crusoe‘s Retreat near Namaqumaqua Village in Serua
Province and Fijian Shangri-La Resort near Sigatoka Town in Nadroga.
These areas represent the continuation of fringing and back reef platform
that typifies this coastline. Attractive reef and beach features combined
with moderate rainfall stimulated rapid development along the Coral
Coast after the sealing of the Queens Highway in the early 1970s
(Thaman, 2002). This has resulted in erosion, habitat loss, elevated
27
siltation, pollution, and the degradation of near shore habitats, such as
mangroves, seagrass beds and coral reefs. Visitors at tourist resorts
make use of the coastline for activities such as scuba diving, snorkelling,
glass bottom boat rides, fishing, kayaking, and sailing. Tourism forms an
integral part of the local economy as a number of local people work in
hotels or related industries (Coral Cay, 2005).
Small settlements and households are scattered along the coastline,
however the majority of the population live in villages of between 100 and
300 people or in the market town Sigatoka, which is central to the region
and houses more than 8000 people. In addition to tourism, the fishing
grounds of the Coral Coast region provide for a large proportion of the
local diet (Coral Cay, 2005).
The southwest coastline of Viti Levu is steeply shelving offshore; the
200m depth contour lies approximately 1km from shore. Fringing reef
extends along the Coral Coast for approximately 63 kilometres and up to
1000 metres offshore. Behind the break zone, back reef habitat extends
over the comparatively flat platform towards shore. The continuity of the
reef is periodically broken by channels cut through the reef due to fresh
water influx from rivers and streams and sediment deposition. These
channels provide suitable habitat for corals, other sessile forms, and
their associated communities below the spring low tide (Coral Cay, 2005).
28
In particular, the two primary wastewater treatment initiatives that were
monitored as part of this study include a subsurface flow constructed
wetland at Tagaqe Village near Hideaway Resort, and a cost effective
commercial wastewater management system at Crusoe‘s Retreat near
Namaqumaqua Village (Figure 12). However, constant water quality
monitoring on general sites along the Coral Coast as a follow up on the
findings of Mosley and Aalbersberg (2003) was also undertaken for other
locations (Figure 13 and Table 1).
Figure 12: Map of Coral Coast showing the location of the two primary wastewater
treatment systems being monitored (monitored systems circled).
29
Figure 13: Location of general Coral Coast sampling sites, villages and hotels
[after Mosley and Aalbersberg, 2003].
Table 1: List of general Coral Coast water quality sampling sites, similar to Mosley
and Aalbersberg (2003)
Site Number Location
1 Fijian Resort – ocean side 2 Outrigger resort-western side 3 Tubakula resort-eastern side 4 West of Navola Village 5 East of Votua Village 6 Tagaqe Village 7 Sovi Bay Beach 8 Hideaway resort-western side 9 Front of Naviti resort 10 West of Komave Village 11 Tabua Sands resort 12 Vatukarasa Bay 13 Malevu Village-eastern side 14 Crows Nest resort 15 Korotogo Bridge 16 Matai Kandavu Beach 17 Between Malevu/Vatukarasa Villages 18 Warwick Hotel
30
2.3.2. Human impacts on the coastal environment
Human impacts on the coastal aquatic systems in Fiji are becoming
evident. Poor sanitation and inadequate waste treatment and disposal
practices are visible in most places, resulting in significant degradation
of coastal ecosystems both in rural and urban centres (Singh, 2001).
In addition, much pollution derived from industrial wastes, agricultural
runoffs, wastewater effluents, and faecal inputs are likely to be trapped
in enclosed, slow flowing, shallow water muddy environments, especially
adjacent to villages, towns, industries, or tourist hotels. The problem is
disastrous when fish stocks that are important for fisheries are depleted
as a result of loss of certain inter-related ecosystems and overfishing.
Shellfish can also assimilate toxic pollutants over a period of time thus
increasing the toxicology effects on humans (Singh, 2001).
In Fiji, particularly along the Coral Coast the principal contributor to
human impacts on coastal environments can be attributed to the
inadequate wastewater treatment and disposal from sources including
tourist resorts and hotels, coastal village households, piggery, and
agricultural runoffs through rivers and creeks. Therefore, strengthening
of legislation and self motivation is paramount to achieve a healthier
coastal environment (Tanner and Gold, 2004).
31
2.4. Water quality standards for normal coral growth
Research on coral reefs in other locations has found that the critical
nutrient levels considered healthy for coral reefs without being overgrown
by algae are approximately 1.0 mol/L of nitrogen (N) as nitrate,
ammonia or nitrite (14 g/L N) and 0.1 mol/L of phosphorus (P) as
orthophosphate and organophosphate (3 g/L P) (Bell et al., 1987; Bell,
1992; Goreau and Thacker, 1994).
However, recent studies (Blake and Johnson, 1988; Brodie et al., 1989)
on Australian nearshore fringing reefs in good condition have found
relatively high nitrogen levels within the range of 1.5 - 2 mol/L. The
required concentration of dissolved phosphates found by Crossland and
Barnes (1983) for normal coral growth was within the range of 0.11 -
0.32 mol/L. However, phosphate levels as high as 0.74 mol/L had
been reported from studies of Australian fringing reefs (Blake and
Johnson, 1988).
For relatively unpolluted open ocean waters, nitrate concentrations
usually range from 0.5 – 4.8 mol/L (cited in Naidu et al., 1991). Nitrite
levels for unpolluted waters often varied from 0 – 0.22 mol/L (Wetzel,
1975). Besides that, ammonia levels are a better indicator for sewage
pollution and anaerobic conditions compared to nitrate and nitrite
(Hawker and Connell, 1992). For the Astrolabe lagoon where there is
32
insignificant pollution, levels of ammonia obtained were in the range 0.05
– 0.20 mol/L. Nitrite concentrations ranged from 0.05 – 0.24 mol/L
(Yamamuro et al., 1991). This lagoon represents a relatively unpolluted
site.
According to Wetzel (1975), total phosphorus levels in unpolluted surface
waters range between 0.32 mol/L and 1.6 mol/L. In freshwater rivers
and creeks, the critical total phosphorus levels fall in the range 1.0 – 3.2
mol/L. With regards to dissolved phosphates, the range of
concentrations in unpolluted natural waters extends over a wide range
from 0.01 mol/L to 2.1 mmol/L in some cases. Values obtained for the
Astrolabe lagoon seagrass bed was lower at 0.08 – 0.15 mol/L
(Yamamuro et al., 1991).
Moreover, in open ocean seawater production is thought of as being
influenced by the mole ratio of concentrations of nitrogen to phosphorus
in the water. For average seawater, the nitrogen (N) to phosphorus (P)
mole ratio is about 15 N: 1 P which reflects the ratio of their utilisation
by phytoplankton (Collier, 1970). For nearshore waters the expected
nitrogen to phosphorus mole ratio was found to be around 10 N: 1 P
(Blake and Johnson, 1988; Mosley and Aalbersberg, 2003).
33
Furthermore, there are currently no country specific water quality
standards in Fiji as most authorities are using the World Health
Organisation (WHO) Drinking Water Quality guidelines (WHO, 2004). But
in Australia and New Zealand, the recommended water quality standards
for nearshore waters to support normal coral growth and for recreation
are summarized in Table 2.
Table 2: Summary of recommended standards for nearshore waters to support coral reefs and recreation in Australia and New Zealand [ANZECC, 2000]
Parameter Coastal water quality standards
pH 8.0 – 8.4 Dissolved oxygen (mg/L) >6 Clarity (m) >1.2 Total nitrogen (µmol/L) <7.14 Total phosphorus (µmol/L) <0.48 Nitrate and nitrite (µmol/L) 0.14 – 0.57 Phosphate (µmol/L) 0.16
Faecal coliform (counts/100ml) <150
2.5. Status of water quality in Fiji and the Coral Coast
A water quality study on unpolluted nearshore waters in Fiji‘s Great
Astrolabe Reef and lagoon has revealed low average nutrient
concentrations approximately 0.74 micromoles/litre (mol/L) of nitrate,
and 0.07 mol/L of phosphate. The nitrogen (N) to phosphorus (P) mole
ratio was around 10 (Morrison et al., 1992).
Another nutrient study of moderate polluted coastal waters along the
Coral Coast in Fiji (Mosley and Aalbersberg, 2003) found that levels for
nitrate and phosphate exceeded thresholds considered harmful to coral
34
reef ecosystems. Furthermore, nutrient levels were highest at sites
located near hotels, other populated coastal locations and in rivers. The
study has yielded nitrate concentrations ranging from 0.1 – 7.01 mol/L
with a mean of 1.69 mol/L for seawater samples. The nitrate values for
river water samples ranged from 1.9 – 24.7 mol/L with a mean of 10.8
mol/L. The phosphate levels for seawater varied between 0.07 – 1.51
mol/L with an average of 0.21 mol/L. For freshwater samples,
phosphate concentrations ranged from 0.50 – 3.40 mol/L with a mean
of 1.30 mol/L. The mole ratio of nitrogen to phosphorus for seawater
samples was 8 whilst for freshwater the mean N: P ratio was 12.
In addition to coastal development, fishing in Fiji, which occurs at both
traditional subsistence and commercial scales, has significantly reduced
the populations of many species. Although data are scarce, even
traditional techniques, such as hand-lines, fish traps and gill nets, in
combination with commercial catches have led to over-fishing of many
reef areas (Coral Cay, 2005). For example, an earlier study by Jennings
and Polunin (1996) found low abundances of certain highly targeted fish
species, such as Groupers and Emperors. Over-fishing of prized
invertebrate species such as Tridacna clams and Sea Cucumbers has
also been reported close to urban areas and is thought to have increased
since the introduction of scuba apparatus and escalating demands of
foreign markets (Vuki et al., 2000).
35
Fiji is the world‘s second largest exporter of live reef products for the
aquarium trade (Wilkinson, 2002) with a well-established industry that
has been operating for over 16 years exporting coral reef fishes and curio
coral (Lovell, 2001). The anthropogenic threats to reef health have been
compounded by natural and semi natural threats such as storm damage,
outbreaks of the coral eating Crown-Of-Thorns starfish (Acanthaster
planci) and coral bleaching events (Coral Cay, 2005). Bleaching events
occur during occasional periods when climate conditions raise seawater
temperatures and solar irradiance and cause a paling of coral tissue from
the loss of symbiotic zooxanthellae (Brown, 1997; Westmacott et al.,
2000).
A major coral bleaching event occurred in Fiji in March and April 2000
and had large-scale effects throughout the country, including the
Mamanucas region. For instance, South and Skelton (2000) reported
bleaching of up to 90 percent of coral colonies with up to 40 percent
mortality (Wilkinson, 2002), although there was significant spatial
variation in its severity throughout Fijian waters. There is evidence that
many of the corals recovered but mortality was certainly significant
although it is difficult to quantify because of the limited long-term
monitoring data available (Coral Cay, 2005). A second less severe
bleaching event occurred in the Mamanucas in April 2002 but did not
36
significantly alter the percentage cover of live hard coral (Walker et al.,
2002).
According to a recent study of the Coral Coast in Fiji (Tanner and Gold,
2004), nitrogen export from the coastal land use practices assessed in
that study have increased by more than 60 percent in the past 20 years.
The study extrapolated that if nitrogen control measures are not adopted,
at the current growth rate of 2.7 percent per annum for village and
tourist populations, nitrogen export in 2014 (e.g. 37,500 kg/year) will be
more than double the 1984 levels of 16,800 kg/year. However with the
adoption of widespread nitrogen control practices and initiatives,
nitrogen exports in 2014 (e.g. 17,500 kg/year) could be comparable to
1984 levels even at the current annual Coral Coast human growth rate of
2.7 percent.
The study indicated that around 60 percent of nitrogen export to the
Coral Coast from key coastal sources excluding rivers and streams can
be derived from village wastewater. Small piggeries account for 30
percent of nitrogen inputs while the remaining 10 percent can be
attributed to tourist resort and hotel wastewater (Tanner and Gold,
2004).
37
Moreover a Japanese International Cooperation Agency (JICA) funded
review of wastewater management by tourist resorts on the Suva, Coral
Coast, Nadi and Mamanuca areas has found that the standard of
wastewater in Fiji resorts is poor. Results showed that only 5 out of 11
sewage treatment plants (STPs) and 3 out of 5 septic tanks assessed have
satisfied recommended levels for biological oxygen demand (BOD). For
total suspended solids, only 3 out of 11 STPs and 4 out of 5 septic tanks
analysed achieved World Health Organization (WHO) recommended
standards. With regards to faecal coliform and total nitrogen, 4 out of 11
STPs complied with recommended standards of 200 counts/100ml
(c/100ml) or less and 714 µmol/L or less, respectively. The
recommended level for total phosphorus in a STP is 32 µmol/L or less.
Unfortunately, only 1 out of 11 STPs sampled has fulfilled this
requirement (Thaman and Sykes, 2005).
Research on the highly populated nearshore waters of Suva Harbour and
Laucala Bay by Naidu & Morrison (1988) and Naidu et al. (1991) showed
that the concentration of nitrogen as nitrate (NO3-N) in the harbour
ranged from <0.7 µmol/L to 2.5 µmol/L. The total phosphorus
concentrations (PO4-P) ranged from 0.19 – 2.2 µmol/L which was
comparable to results by Campbell et al., (1982).
38
In general, clarity of the Suva Port waters in 1987 – 1988 ranged from
0.5m to 5m with a mean of 1.7m (Naidu et al., 1991), which has
decreased significantly from the observed 3.1m in 1982 (Campbell et al.).
Temperature variations within the water column were less than 0.5 C
and seldom exceeded 1 C. Salinity of surface water in the Laucala Bay
and Suva Harbour varied from 7 parts per thousand (ppt) to 35 ppt.
Respective mean nitrate and phosphate levels of 17.29 µmol/L and 5.61
µmol/L were also attained for the Suva Port nearshore waters by Naidu
et al., (1991). Dissolved oxygen level in the Suva Port in 2003 ranged
from 1.8 mg/L to 8.2 mg/L with a mean of 6.1 mg/L (MS312 Class,
unpublished Report).
Campbell et al. (1982) found extremely high faecal coliform levels from
various rivers discharging into Laucala Bay, especially during periods of
heavy rainfall when septic tank effluent and sewage effluent from pit
latrines seep into creeks and rivers. In 1981 when population was not as
high as today, sewage population was already significant. Sanitary
surveys by Corless (1995) and a Marine Pollution (MS312) Class in 2003
also found alarming faecal pollution in the Suva nearshore area (Table 3).
39
Table 3: Comparison of the mean faecal coliform counts around Suva nearshore
waters and rivers from different studies (cfu/100ml)
Site Campbell et al., (1982) Corless (1995) MS312 Class 2003
Vatuwaqa River 5,100 20,100 14,900 Raiwaqa Outfall 350,000 not assessed not assessed Samabula River 830 not assessed not assessed Nasinu River 4,200 not assessed not assessed Rewa River 110 not assessed not assessed Nabukalou Creek not assessed 6,267 10,000 Walu Bay bridge not assessed 670 >20,000 Tamavua River not assessed 3,700 9,400 Navesi River not assessed 1,500 1,900 Vatuwaqa I/estate not assessed 10 29,000 Nasese foreshore not assessed 100 17,600 MSP jetty not assessed 23 110
Kings wharf not assessed 140 1,700
Another study of the Port of Suva by Tamata et al. (1992) found variable
results for different water quality parameters as summarised in Table 4.
Table 4: Water quality results for the Port of Suva in 1992 as observed by Tamata
et al.
Parameter Range Mean
Temperature (C) 22 – 28.5 27.7 pH 7- 8.5 8.0 Salinity (ppt) 0.5 – 35 27.6 Clarity (m) 0 – 3 1.9 Dissolved oxygen (mg/L) 3.2 – 9.8 6.0 Total Kjeldhal nitrogen (µmol/L) 3.57 – 271.4 60.7 Nitrates (µmol/L) 0 – 98.57 5.47 Nitrite (µmol/L) 0 – 6.02 0.59
Ammonia (µmol/L) 0 – 184.4 6.61
Total phosphorus (µmol/L) 0.1 – 14.2 1.63 Phosphate (µmol/L) 0 – 10.32 1.26 Faecal coliform (counts/100ml) 0 – 8.5 x 106 156,917
Furthermore an assessment of nutrients in Laucala Bay from 2003 –
2004 found nitrate values between 0.87 µmol/L and 27.05 µmol/L with a
mean of 1.77 µmol/L. Phosphate varied between 0.46 – 11.01 µmol/L
with a mean of 0.95 µmol/L (Singh & Mosley, unpublished). Another
survey in 2004 yielded mean nitrate and phosphate levels of 3.68 µmol/L
and 1.20 µmol/L, respectively (Taloiburi, unpublished).
40
Chapter 3 Nutrient Enhancement and Coastal Waters
3.1. Introduction
This chapter outlines the potential effects of nutrient enrichment on
coral reefs, as observed from previous case studies elsewhere. This is of
uttermost importance because it highlights the deleterious implications
that coral reef ecosystems can face if proper wastewater management
practices and initiatives are not adopted. Sections covered include
importance of coastal aquatic systems, potential effects of nutrient
enhancement on coral reefs with proven case studies of the Kaneohe Bay
in Hawaii, the Florida Keys, and the Discovery Bay in Jamaica.
3.2. Importance of coastal aquatic systems
Coastal aquatic systems, including estuarine and marine nearshore
environments, deserve human attention for three primary reasons. First,
healthy coastal systems serve as shelter and food for numerous plants
and animals that are vital for biodiversity. Another reason for
maintaining coastal environments can be attributed to aquatic systems
being utilized for commercial and recreational activities. The third
implication is that any coastal or inland activities can pose threats to the
health of coastal ecosystems (Hauxwell et al., 2001).
A distinguishable difference between a coastal system and an offshore
oceanic environment is that a coastal zone is where inputs of nutrients
and other materials from the land are a key feature, whilst an offshore
41
boundary begins from a zone that receives insignificant influence from
land inputs. Consistent loading from land activities into coastal
environments tends to support rapid growth and reproduction of primary
producers and consumers, making these areas the most highly
productive in the world (Taylor et al., 1995). Research shows that
although coastal waters represent only 10 percent of the total ocean
surface, they account for 20 percent of total production and 50 percent
of total fish production in the oceans (Ryther, 1969).
Figure 14: Mangroves are a classic example of a coastal aquatic system
In general, coastal aquatic ecosystems (Figure 14) including coral reefs,
mangroves and seagrass beds are directly degraded by natural
disturbances (Short and Echeverria, 1996) and human activities (Sargent
et al., 1995) even though the linkage process may involve multiple steps
(Kemp et al., 1983). But it is the indirect effects of excess nutrient
addition from watersheds to coastal waters that are cited as the most
42
pervasive human impacts on coastal areas and coral reefs (GESAMP,
1990; Short et al., 1993; National Research Council, 1994; Valiela et al.,
1997).
Of all the essential nutrients, nitrogen and phosphorus are the two key
nutrients that often limit the growth of primary producers. Hence, excess
addition to coastal waters will yield genuine concerns, as they are likely
to influence the survival of corals. In freshwater environments,
phosphorus is often the limiting nutrient implying that the addition of
phosphorus stimulates primary productivity. Nitrogen is often limiting in
marine environments, which means that loading of nitrogen will boost
growth and reproduction of primary producers (Hauxwell et al., 2001).
However, it is difficult to accurately predict the amount of nutrients that
can be safely added to coastal waters, despite studies showing the
existence of links between nutrient supply, algal production, and
degradation of coral reefs (Lapointe and Clark, 1992; Goreau and
Thacker, 1994; McCook, 1999; Szmant, 2002). Thus, it is essential for
scientists, governments, non government bodies, managers and
community members to work together to develop research initiatives and
monitoring programs that detect and predict small, relevant changes
caused by increased nutrient loads (Hauxwell et al., 2001).
43
3.3. Potential effects of nutrient enrichment on coral reefs
Coral reefs are highly productive with a considerable biodiversity (Goreau
and Thacker, 1994). They are extremely important in the Pacific for
fisheries, tourism, shoreline protection, source of protein, medicine,
employment, and income for many coastal people (Mosley and
Aalbersberg, 2003).
Despite coral reefs being more productive and species rich, they are very
sensitive to elevated levels of nutrients. Increased development of the
coastline and utilisation of coastal resources over the past years have
contributed to significant degradation of reef habitats and loss of species
diversity (Hodgson, 1999).
Maintaining the health of coral reefs within the South Pacific is therefore
necessary in protecting coastal infrastructure (seawalls, wharves, roads,
houses and hotels) and employment (in fisheries, tourism and services).
In recent years, several reef systems throughout the Pacific have
significantly deteriorated compromising the valuable ecological services
that reefs provide (Goreau and Thacker, 1994). Degraded reefs have most
corals replaced by fleshy algae thus, supporting only a limited fish
population and virtually lacking both growing corals, which break waves
in shallow water and sand producing algae such as Halimeda that helps
44
to renourish a small fraction of beaches (Bell, 1992). These impacts have
been attributed to various factors such as eutrophication (increased
growth of algae due to elevated nutrient levels in water), human physical,
increased erosion on land and siltation of reefs, hurricanes, overfishing,
climate change, and water pollution.
Nutrients such as nitrogen and phosphorus are required for the growth
of phytoplankton and other algae, which form the base of the ocean food
chain. Inorganic nitrogen exists predominantly as ammonia, nitrate and
nitrite whilst phosphorus is present as orthophosphate and
organophosphate. These two nutrients exist in seawater naturally.
However, when levels exceed concentrations considered healthy for coral
survival, reefs may be overgrown by weedy macroalgae resulting in
deleterious effects. The problem is severe in tropical waters because
nutrient concentrations capable of damaging corals are lower than
temperate regions (Naidu et al., 1991; Mosley and Aalbersberg, 2003).
Increased nutrient loading into coastal waters from sewage outfalls,
household wastes, fertilised farmland, chemicals, cleaning detergents,
and industrial effluents may lead to an abundant content of organic
material which is gradually decomposed by micro-organisms in the
water. This process enables algal growth, which eventually results in
eutrophication within the waters receiving emissions. The process of
45
decomposition consumes oxygen and sometimes can lead to oxygen
deficiency and fish kills. Nevertheless, this impact is limited nowadays
due to improved sewage treatment and disposal (Goreau and Thacker,
1994; McCook, 1999).
On the other hand, effects of moderate eutrophication of originally
nutrient poor water are not entirely negative. Increased growth of algae
and other vegetation can be beneficial to the aquatic fauna, at least to
begin with. Fish production rises, for instance (Clark, 2002). But if
eutrophication continues, plankton growth becomes so great that it
eventually clouds the water. The resultant darkness below the surface
can be harmful to benthic vegetation. Such a process favours algae and
plankton eating fish including parrotfish (Scarids), surgeonfish,
damselfish (Pomacentrids), and unicorn fish (Acanthurids), while
depleting the numbers of predator fish species more sought after for
human consumption and commercial purposes (Goreau and Thacker,
1994).
In highly nutrient rich waters, plankton production can be copious
indeed. Certain plankton species appear intermittently in massive
quantities, in what is termed ―algal bloom‖. Such algae can give the water
an unpleasant smell or taste and some are even poisonous. The best-
documented examples of algal bloom were from the Hong Kong Harbour
46
(Songhui Lu and Hodgkiss, 2004) and Seto Sea in Japan (Imai and Itoh,
1987; Fisheries Agency, 2000). In these places harmful algal blooms
steadily became prevalent as the human population and water pollution
increased. The linkage between water pollution and algal bloom was
evident in the Japanese experience where harmful algal blooms suddenly
declined after pollution control measures were introduced (Castro and
Huber, 2003).
Recent research on the shift from coral to macroalgae dominance (a
common picture in reefs experiencing eutrophication) indicates that this
effect, presumed to be a direct impact from liquid nutrient enhancement,
may not be exactly as imagined. Apparently, accomplishing the shift from
coral to macroalgae dominance also requires that population of
herbivorous fish be significantly depleted (Ginsburg, 1994; McCook,
1999). Therefore, the influence of other factors such as the abundance of
herbivores (e.g. sea urchins, grazing fishes) to graze the algae is equally
important (Aronson and Precht, 2000). Overfishing does contribute to the
shift from coral to algal dominance and not primarily driven by nutrient
enhancement alone (Neckles et al., 1993).
Nevertheless, nutrient enrichment alone has the potential to harm
corals. According to Olivieri (1997), eutrophication also operates at the
zooxanthellae level. Excess nutrients increase zooxanthellae growth,
47
which, counter intuitively, is not beneficial for the coral host.
Zooxanthellae populations under natural conditions are constant and
nutrient limited, particularly by nitrogen within the coral host. However
with excess nutrients the zooxanthellae population grows uncontrolled
and the balance of nitrogen-carbon fluxes between the coral host and
zooxanthellae is abruptly disrupted resulting in reduction and weakening
of coral calcareous skeletons.
In addition, nutrient enrichment increases coral diseases such as coral
bleaching, white pox and the black band disease. It also enables the
growth of animal competitors like filter feeding sponges, polycheate
worms, boring molluscs and ascidians, many of which bore into corals
and weaken their skeletons as well as displacing corals and accelerating
the bioerosion of reefs. The problem is worse in areas where there is
overfishing (Goreau and Thacker, 1994; McCook, 1999; Mosley and
Aalbersberg, 2003).
Moreover, the overgrowth of macroalgae as an impact of nutrient
enrichment can lead to mortality and loss of biodiversity of live corals
and a loss of settlement sites for coral larvae. Overgrowth of algae may
also impact fish and invertebrate biodiversity due to the resulting habitat
homogeneity (McCook, 1999).
48
According to Thrash (2003), nutrient enhancement in tropical waters has
the most significant effect on reef building corals. Specifically inorganic
forms nitrogen and phosphorus have the greatest effect on coral
mortality and reproduction. Production of viable gametes and successful
fertilisation has been found to reduce as a consequence of excess
nutrients. Phosphorus itself has proven to dramatically reduce
fertilisation and stimulate growth of irregular embryos. Corals exposed to
elevated amounts of ammonium produced smaller and fewer eggs and
had less testes material compared to unexposed corals. These effects
appear if the nitrogen level exceeds 1.0 mol/L and phosphorus
concentration above 0.1 mol/L (Thrash, 2003).
Furthermore, high levels of phosphorus can lead to reduction in the
structural density of stony corals, causing them to lose their strength
and crumble (Mosley and Aalbersberg, 2003).
Currently, some of the best studied areas of the effects of nutrient
enrichment on coral reefs are in the Kaneohe Bay in Hawaii (McCook,
1999; Szmant, 2002), Florida Keys (Lapointe and Clark, 1992; Thrash,
2003) and the Discovery Bay in Jamaica (Goreau, 1992; Goreau and
Thacker, 1994). The incidents showed that elevated nutrient levels favour
growth of planktonic algae and large macroalgae, which would normally
grow slowly in low nutrient waters. Thick algal turfs are likely to smother
49
corals. The problem is disastrous in areas where there is overfishing
since grazers like parrotfish are depleted. Grazer fish usually controls
algal growth from overshadowing and replacing corals (Goreau and
Thacker, 1994).
3.3.1. Kaneohe Bay case study
Kaneohe Bay, located on the northeast shore of the island on Oahu, once
had some of the most luxuriant reefs in Hawaii. Until the 1930s the area
around the bay was sparsely populated. In the years leading up to the
World War Two, with military build up of Oahu, the population began to
rise. The increase continued after the war as the shores of the bay were
developed for residential use (McCook, 1999; Szmant, 2002; Castro and
Huber, 2003).
The sewage from this expanding population was dumped right into the
bay. By 1978 about 20,000 cubic metres (over 5 million gallons) of
sewage were dumped into the bay every day. Long before then, by the
mid 1960s, marine biologists began to notice disturbing changes in the
middle of the bay. Loaded with nutrients, the sewage acted as fertiliser
for seaweeds. The green bubble algae (Dictyosphaeria cavernosa) found
the conditions agreeable and grew at an alarming rate, literally covering
the bottom in many parts of the bay. Bubble algae began to overgrow and
smother corals. Phytoplankton also multiplied with the increase in
nutrients, clouding the water. Kaneohe Bay‘s reefs began to die because
50
of such accelerated algal growth due to excess nutrient loading (Castro
and Huber, 2003).
In 1978 public pressure managed to help in the great reduction of
sewage discharge into the bay, as sewage was diverted offshore. The
result was dramatic. Bubble algae in much of the bay died and corals
began to recover at an unexpected fast rate. By early 1980s, bubble algae
were fairly scarce and corals had started to grow healthier. The reefs
were not what they once were, but they seemed to be on track for
recovery (Szmant, 2002; Castro and Huber, 2003).
However in November 1982 Hurricane Iwa struck Kaneohe Bay. During
the years of pollution a layer of coral skeleton had weakened, becoming
fragile and crumbly. During the hurricane, this weak layer collapsed and
much of the reefs were severely damaged. Fortunately corals were
already recovering enabling broken pieces to grow back (Goreau and
Thacker, 1994; Castro and Huber, 2003).
The rapid recovery of reefs in Kaneohe Bay observed during the early
1980s did not continue. By 1990 the recovery seemed to have leveled off.
Some coral areas began to decline allowing bubble algae to become
abundant. There were a number of possible explanations for this. Even
though most sewage was now discharged outside the bay, some
51
nutrients continue to enter the bay from boats, septic tanks and
cesspools of private homes and other sources. Besides that, nutrients
from old sewage outfalls adsorbed and accumulated onto sediments were
being slowly released even after the outfall was diverted offshore. There
was also evidence that overfishing had reduced grazing fish population
that was likely to graze on the bubble algae, thus controlling their growth
(Szmant, 2002; Castro and Huber, 2003).
In addition, few grazing fish species that remained prefer to eat other
introduced species of seaweed from outside Hawaii rather than the green
bubble algae. Therefore, it appeared to indicate that reefs would continue
to face stress if developments on land are not properly planned and
incorporated with other inter-dependent systems (Castro and Huber,
2003).
3.3.2. Florida Keys case study
Florida Keys is another well researched classic example of the effects of
nutrient enhancement on coral contamination. The Florida Keys are a
chain of approximately 800 independent islands located in Monroe
County off the southeastern tip of Florida, representing the most
southerly point of the continental United States. The Florida Reef Tract is
the most widespread living coral reef system in North American waters
and the third largest system in the world. Extending over 1,550 square
kilometres across southern Florida and the Florida Key archipelago,
52
these reefs consist of a series of ridges and channels that form parallel to
the Straits of Florida. The reefs comprise a bank reef system of almost
continuous reef communities in line that run parallel to each other
(Thrash, 2003).
Unfortunately there are numerous threats to the marine environment in
south Florida and the Keys. Past research has shown that there is
decline of healthy corals; invasion by algae into sea grass beds and reefs;
decline in certain fisheries; increase of coral diseases; and coral
bleaching. Although the cause of these problems, whether natural or
anthropogenic, can be debated, studies showed that land use and
resource exploitation by humans have implicated coral communities in
the Florida Reef Tract (Hauxwell et al., 2001; Thrash, 2003).
On average, over three million visitors and 80,000 fulltime residents
inhabit the Florida Keys each year and there are considerable direct and
indirect effects from land use and utilisation of nearshore systems.
Although sedimentation is the most widespread problem, nutrient
loading of coastal waters is perhaps the most common human impact
plaguing several stretches of the Florida Reef Tract, particularly in the
Florida Keys National Marine Sanctuary. Researchers have noted a loss
of biodiversity in corals and rise in diseased and damaged corals
throughout the National Marine Sanctuary. Some of diseases include
53
white pox and black band disease along with coral bleaching and
macroalgal blooms (Lapointe and Clark, 1992; Thrash, 2003).
The most significant source of nutrient input in the Florida Reef Tract
originated from improperly treated and mismanaged wastewater. In the
Florida Keys alone there were roughly 200 sewage treatment plants,
22,000 septic tanks, 5,000 cesspools and 139 marinas harbouring over
150,000 boats. Nutrients such as nitrate and phosphate are intimately
associated with sewage and carried through the region by more than 700
canals and channels (Thrash, 2003).
Organic nitrogen is carried in ground water reservoirs and is problematic
through seepage of porous limestone. The impacts from deficient
wastewater treatment systems in the Keys extend beyond disease and
contamination of corals. In fact the effects of this deficiency are
beginning to adversely affect water resources and ultimately human
health and welfare of people in southern Florida and the Keys. Therefore,
wastewater management efforts in the Florida Keys need to focus on
elimination of diffuse sources of pollution, particularly nutrients and
faecal loadings (Lapointe and Clark, 1992; Thrash, 2003).
3.3.3. Discovery Bay case study
The third known case of the implications of excess nutrient on coral reef
survival was observed in the Discovery Bay in Jamaica. Coral reefs are
54
the most important natural resource for Jamaica and other Caribbean
islands, providing the bulk of fisheries, marine biodiversity and tourism
returns. However, most parts of Jamaican reefs have significantly
deteriorated in recent years due to enriched nutrient inputs (Goreau,
1992; Goreau and Thacker, 1994).
Nutrients enter the Jamaican coastal environment from streams, creeks,
and submarine springs supplied by groundwater seepage. Measurements
in 1980 within the Discovery Bay found nitrate levels within the range of
5 – 10 µmol/L. By the late 1980s these had risen to around 10 – 15
µmol/L and ecological replacement of corals by weedy algae was nearly
complete. Growth of human population and tourism along the shore in
the 1980s provided local phosphorus inputs which had been previously
lacking, causing rapid eutrophication (Goreau, 1992; Goreau and
Thacker, 1994).
Negril, located at the western tip of the island, had explosive tourism
development and population growth. As a result, Negril was subjected to
unprecedented algae overgrowth that covered the bottom exceeding live
corals (Goreau and Thacker, 1994).
Recent observations of the increasing abundance and species diversity of
algae around Jamaica suggested that eutrophication has become a
55
general phenomenon. Eutrophication has been so severe that many reefs
which formerly had more than 95 percent live coral cover are now more
than 95 percent algal covered (Goreau, 1992). Studies undertaken in the
1990s showed that only least developed and populated areas have coral
reefs in good condition with algae cover of 20 percent or less.
Eutrophication was visible in all populated bays including the Port
Antonio which had the lowest population density due to mountainous
topography and high rainfall. This implies that nutrient concentrations
need to be reduced by 90 – 95 percent to allow ecosystem recovery
(Goreau and Thacker, 1994).
56
Chapter 4 Monitored Wastewater Treatment Systems
4.1. Introduction
This chapter aims to explain the background, design, operating
procedures, and expected performance of the two major cost-effective
wastewater treatment initiatives implemented along the Coral Coast of
Fiji, which had been consistently monitored by this study since 2005.
The two principal wastewater management initiatives that are discussed
in this chapter include a constructed subsurface wetland at Tagaqe
Village and an AdvanTex wastewater treatment system at Crusoe‘s
Resort, both of which are situated along the Coral Coast of Fiji. This
chapter also discusses a model greywater treatment drum system
experiment ex-situ at the Faculty of Islands and Oceans‘ laboratory to
mimic an in-situ system at Votua Village along the Coral Coast.
4.2. Constructed wetland for wastewater treatment
4.2.1. Status of wetlands
Natural wetlands are land areas that have prolonged high water tables or
are at least covered with shallow water which tend to support plants
specially adapted to grow in alternating wet and dry conditions. Classic
examples of natural wetlands include swamps, bogs, sloughs, fens and
marshes (Moore, 1993). A ―constructed wetland‖ is defined as a wetland
specifically constructed for the purpose of pollution control and waste
management, at a location other than existing natural wetlands.
Specifically a constructed wetland is a shallow basin filled with some sort
57
of substrate, usually soil or gravel, and planted with vegetation tolerant
of saturated conditions. Wastewater is introduced at one end and flows
over the surface or through the substrate, and is discharged at the other
end through a weir or other structure which controls the depth of the
water in the wetland (US EPA, 1993).
Historically, natural wetland areas have not made good farmland. As a
result, wetlands were viewed negatively and systematically reclaimed for
alternative development in various parts of the world. Research, however,
showed that wetlands are highly productive ecosystems that support
vigorous plant growth and a broad variety of animals. Wetlands also
improve the quality of water that flows through them by filtering out
impurities, actively degrading waste matter and removing certain
chemicals that flows from upstream. The discovery of this attribute led to
the idea of intentionally using wetlands to treat wastewater (Moore,
1993).
Wetlands incorporate physical, biological, and chemical processes to
treat wastewater. Water flows in and slows down as it spreads across the
wetland surface. This slowing of flow allows soil and sediment particles
to filter or be physically adsorbed. The process also removes nutrients
such as phosphorus and chemicals that are attached to sediments.
58
Biological and chemical treatment processes transform materials
contrary to mere physical removal (Moore, 1993).
Figure 15: A systematic diagram of a horizontal flow constructed wetland [Fujita, 1998]
In addition, constructed wetlands maximise wastewater treatment by
ensuring slow flow rates and extra surface area provided by wetlands
(Figure 15). Plant stems and roots provide surface areas that promote
micro-organism presence, which utilise some of the nutrients and
organic matter carried in runoff water (Smil, 2000).
Constructed wetlands have been enthusiastically adopted in many
societies in other countries including New Zealand, Australia, United
States of America, Canada, Denmark, Czech Republic and so forth as a
cost effective means of efficient secondary and tertiary wastewater
management (Fujita, 1998; Tanner and Sukias, 2002).
However the constructed wetland at Tagaqe Village on the Coral Coast of
Fiji is the first of its kind to be trialled in Fiji or any other small Pacific
Island country.
Inflow Outflow
59
There are two main categories of constructed wetlands comprising (i)
surface flow, and (ii) subsurface flow designs. In a surface flow wetland
(Figure 16), wastewater flows through a shallow pond planted with
emergent plants such as bulrushes, reeds or sedges.
Figure 16: Diagram of a “surface flow” constructed wetland [Kadlec et al., 1996]
In subsurface or gravel bed designs (Figure 17), the wetland if filled with
gravel or similar substrate and plants are grown rooted in the gravel.
Despite that a number of subsurface wetlands were just bare with no
plants at all on the gravel (Tanner and Sukias, 2002).
Figure 17: Diagram of a “subsurface flow” wetland [Kadlec et al., 1996]
4.2.2. How constructed wetlands improve water quality
A constructed wetland is a complex assemblage of water, substrate,
plants (vascular and algae), litter (primarily fallen plant material),
invertebrates (mostly insect larvae and worms) and an array of micro-
organisms (most importantly bacteria). The mechanisms that are
available to improve water quality are therefore numerous and often
interrelated. These mechanisms include: (i) settling of suspended
particulate matter; (ii) filtration and chemical precipitation through
60
contact of the water with the substrate and litter; (iii) chemical
transformation; (iv) adsorption and ion exchange on the surfaces of
plants, substrate, sediment, and litter; (v) breakdown and transformation
of pollutants by micro-organisms and plants; (vi) uptake and
transformation of nutrients by micro-organisms and plants; and (vii)
predation and natural die-off of pathogens (US EPA, 2000).
4.2.3. Nitrogen transformation processes in wetlands
In the biosphere, nitrogen is continuously transformed between organic,
soluble inorganic and gaseous nitrogen forms. The nitrogen cycle is very
complex, and it is hard to control even the most basic transformations
within a wetland. How much nitrogen is removed (i.e. transformed or
removed from the water phase) in a wetland and which nitrogen
processes or nitrogen fluxes that are the most important depend on
water chemistry and other wetland conditions, such as climate,
vegetation, water depth and water flow (Tanner, 2001a; Bastviken, 2006).
The dominant forms of nitrogen in wetlands that are of importance to
wastewater treatment include organic nitrogen, ammonia, ammonium,
nitrate, nitrite, and nitrogen gases. Inorganic forms are essential to plant
growth in aquatic systems but if scarce can limit or control plant
productivity (Mitsch and Gosselink, 1993).
‗Ammonification‘ is the microbial mineralisation of organic nitrogen to
ammonium. This process can be caused by heterotrophic bacteria and
61
fungi (Patrick and Reddy, 1976; Mitsch and Gosselink, 1986). The
formation of ammonia (NH3) occurs via the mineralisation or
ammonification of organic matter under either anaerobic or aerobic
conditions (Keeney, 1973). The ammonium ion (NH4+) is the primary form
of mineralised nitrogen in most flooded wetland soils. The formation of
this ion occurs when ammonia combines with water as follows:
NH3 + H2O NH4+ + OH-
Upon formation, the ammonium ion can be absorbed by the plants and
algae and converted back into organic matter, or the ammonium ion can
be immobilised onto negatively charged soil particles (Mitsch and
Gosselink, 1986).
Ammonium is then transformed to nitrate by the bacterial process
‗nitrification‘. Wetzel (1983) defines nitrification as the ―biological
conversion of organic and inorganic nitrogenous compounds from a
reduced state to a more oxidised state‖. Nitrification is strictly an aerobic
process in which the end product is nitrate (NO3-); this process is limited
when anaerobic conditions prevail (Patrick and Reddy, 1976). The
process of nitrification (i) oxidises ammonium (from the sediment) to
nitrite (NO2-), and then (ii) nitrite is oxidised to nitrate (NO3
-). The overall
nitrification reactions are as follows:
i) 2 NH4+ + 3 O2 4 H++ 2 H2O + 2 NO2
-
62
ii) 2 NO2- + O2 2 NO3
-…………(Davies and Hart, 1990)
Two different bacteria are required to complete this oxidation of
ammonium to nitrate. Nitrosomonas bacteria oxidises ammonium to
nitrite via reaction (i) and Nitrobacter bacteria oxidises nitrite to nitrate
via reaction (ii) (Keeney, 1973).
Nitrate or nitrite is finally reduced to gaseous end products, nitrous gas
and dinitrogen gas, through the bacterial process ‗denitrification‘
(Bastviken, 2006). Nitrate can also be transformed to ammonium during
low redox conditions (Prescott et al., 1990; Vymazal, 2001). According to
Wetzel (1983) ―Denitrification by bacteria is the biochemical reduction of
oxidised nitrogen anions, nitrate-N and nitrite-N, with concomitant
oxidation of organic matter‖. The general sequence as given by Wetzel
(1983) is as follows:
NO3- ---> NO2
- ---> N2O ---> N2
The end products, N2O and N2, are gases that re-enter the atmosphere.
Denitrification occurs intensely in anaerobic environments but will also
occur in aerobic conditions (Bandurski, 1965). A deficiency of oxygen
causes certain bacteria to use nitrate in place of oxygen as an electron
acceptor for the reduction of organic matter (Patrick and Reddy, 1976).
The process of denitrification is restricted to a narrow zone in the
sediment immediately below the aerobic-anaerobic soil interface (Mitsch
and Gosselink, 1986; Nielson et al., 1990). Denitrification is considered
63
to be the predominant microbial process that modifies the chemical
composition of nitrogen in a wetland system and the major process
whereby elemental nitrogen is returned to the atmosphere (Patrick and
Reddy, 1976; Richardson et al., 1978; Johnston, 1991; Vymazal, 2001;
Trepel and Palmeri, 2002). To summarise, the nitrogen cycle is completed
as follows: ammonia in water, at or near neutral pH is converted to
ammonium ions; the aerobic bacterium oxidises ammonium to nitrite;
then nitrite is converted to nitrate. Under anaerobic conditions, nitrate is
reduced to relatively harmless nitrogen gas that is given off to the
atmosphere (Figure 18).
Figure 18: A simplified diagram of the nitrogen processes and the flows of
different nitrogen forms in a wetland [adapted from Bastviken, 2006]. “ON” is organic nitrogen, “AN” ammonium nitrogen and “NN” nitrate nitrogen.
64
Nitrogen fixation is a bacterial process, which transfers dinitrogen gas to
ammonium. However, nitrogen fixation is generally not significant in
nitrogen rich waters such as constructed wetlands for water treatment
(Kadlec and Knight, 1996). Nitrogen assimilation is the transformation of
inorganic nitrogen to organic nitrogen in cells and tissue. Plant
assimilation only accounts for a small percent of the total nitrogen
removal when the nitrogen load is high (Tanner et al., 1995), which is the
case in most free water surface treatment wetlands (Kadlec and Knight,
1996).
Furthermore, plant uptake does not play an important role in the annual
nitrogen removal in the long run as the nitrogen assimilated by
vegetation usually is released during decomposition of litter (Johnston,
1991). However, plant uptake can contribute to the seasonal dynamic of
nitrogen removal and account for a significant part of the wetland
nitrogen removal during the growth period with rapid nitrogen uptake
(Bastviken, 2006).
Ammonia volatilisation is a physicochemical process where ammonia in
the ammonium-ammonia equilibrium is transported to the gas phase.
Ammonia volatilisation can be important in wetlands with high
temperature and pH, although volatilisation losses of ammonia usually
are small if pH is below 8 (Reddy and Patrick, 1984). Other nitrogen
65
fluxes that can occur in a wetland are sedimentation and resuspension of
the various nitrogen forms. Burial of organic nitrogen in the sediment
will make the nitrogen less available to living plants and organisms,
while release of nitrogen from biomass during decomposition will make
the nutrients available again. Ammonium can also readily adsorb to
sediment particles and litter, ammonium adsorption, as ammonium is a
positively charged ion. The adsorbed ammonium concentration is in
equilibrium with the ammonium concentration in the water, and can be
released if a change in the water chemistry or hydrology occurs
(Bastviken, 2006).
Nitrification and denitrification are influenced by factors at different
spatial scales. In Figure 19, factors affecting denitrification have been
described at different spatial scales; the process scale, the wetland scale
and the landscape scale. At the process scale, the process rates
performed by the bacteria are directly regulated by factors like
temperature and redox conditions. These process scale factors can be
affected by factors on the wetland scale, such as water flow, nutrient
load and different plant communities. Finally, the wetland scale factors
are affected by landscape factors such as climate and land use (e.g.
higher nitrogen load the more agricultural areas there is in the upstream
catchment area) (Bastviken, 2006).
66
Figure 19: Factors affecting the biological processes of denitrification on different
spatial scales [adapted from Trepel, 2002].
On the wetland scale, plants can play an important role for the total
removal of nitrogen in wetlands. Wetlands containing plants have been
shown to remove larger quantities of nitrate than unplanted wetlands
(Tanner et al., 1995; Zhu and Sikora, 1995; Bachand and Horne, 2000;
Lin et al., 2002). However planted systems only showed small
improvements in disinfection, Biological Oxygen Demand (BOD),
Chemical Oxygen Demand (COD), and suspended solids removal. Direct
nutrient uptake by plants was insufficient to account for more than a
fraction of the improved removal shown by planted systems (Tanner,
2001).
67
4.2.4. Phosphorus transformation processes in wetlands
Phosphorus occurs in natural waters and wastewater primarily as
phosphates. They are classified as orthophosphates, and organically
bound phosphates. They may be in solution or particulate form. Organic
phosphates are formed primarily by biological processes and are found in
raw wastewater as food residues and body wastes, and in treated
wastewater as living or nonliving biota (e.g. algae and bacteria from
treatment ponds). Inorganic phosphorus found in wastewater most often
comes from various forms of personal and commercial cleaning solutions
or from the treatment of boiler waters. Storm waters carry inorganic
forms of phosphorus from fertilizers into combined sewers (US EPA,
2000).
Dissolved organic phosphate and insoluble inorganic and organic
phosphate are not usually available to plants until transformed to a
soluble inorganic form. These transformations may take place in the
water column by way of suspended microbes and in the bio-films on the
emergent plant surfaces and in the sediments. Uptake of phosphates by
micro-organisms including bacteria and algae acts as a short-term,
rapid-cycling mechanism for soluble and insoluble forms. Cycling
through the growth, death, and decomposition process returns most of
the phosphate back into the water column. Uptake by the macrophytes
occurs in the sediment pore water by the plant root system. Uptake
68
occurs during the growth phase of the plant and release occurs during
plant death followed by decomposition in the plant litter (US EPA, 2000).
The removal and storage of phosphorus from wastewater can only occur
within the constructed wetland itself. According to Mitsch and Gosselink
(1986) phosphorus may be sequestered within a wetland system by the
following: (i) The binding of phosphorus in organic matter as a result of
incorporation into living biomass, and (ii) precipitation of insoluble
phosphates with ferric iron, calcium, and aluminum found in wetland
soils.
4.2.5. Advantages of constructed wetlands
Constructed wetlands are a cost-effective and technically feasible
approach to treating wastewater and runoff for several reasons: (i)
wetlands can be less expensive to build than other treatment options; (ii)
operation and maintenance expenses (energy and supplies) are low; (iii)
operation and maintenance require only periodic, rather than
continuous, on-site labour; (iv) wetlands are able to tolerate fluctuations
in flow; (v) they facilitate water reuse and recycling; (vi) they provide
habitat for many wetland organisms; (vii) they can be built to fit
harmoniously into the landscape; (viii) they provide numerous benefits in
addition to water quality improvement, such as wildlife habitat and the
aesthetic enhancement of open spaces; and (ix) they are an
environmentally sensitive approach (US EPA, 2000).
69
4.2.6. Limitations of constructed wetlands
There are limitations associated with the use of constructed wetlands:
They generally require larger land areas than do conventional
wastewater treatment systems. Wetland treatment may be
economical relative to other options only where land is available
and affordable.
Performance may be less consistent than in conventional
treatment. Wetland treatment efficiencies may vary seasonally in
response to changing environmental conditions, including rainfall
and drought. While the average performance over the year may be
acceptable, wetland treatment cannot be relied upon if effluent
quality must meet stringent discharge standards at all times.
The biological components are sensitive to toxic chemicals, such as
ammonia and pesticides.
Flushes of pollutants or surges in water flow may temporarily
reduce treatment effectiveness.
They require a minimum amount of water if they are to survive.
While wetlands can tolerate temporary draw-downs, they cannot
withstand complete drying.
The use of constructed wetlands for wastewater treatment and
stormwater control is a fairly recent development. There is yet no
consensus on the optimal design of wetland systems nor is there
much information on their long-term performance (US EPA, 2000).
70
4.2.7. Description of the constructed wetland at Tagaqe Village
The constructed wetland at Tagaqe Village on the Coral Coast of Fiji is a
―subsurface flow‖ gravel bed design (Figures 20 & 21). It was set up in
December 2004 by villagers and staff from the University of the South
Pacific under principal supervision of a scientist from the National
Institute of Water and Atmospheric Research (NIWA) in New Zealand. The
wetland at Tagaqe is a large rectangular hole about 11 m long, 4 m wide
and 0.4 m deep lined with heavy duty plastic and filled with 19 m3
gravel. It was estimated for 15 people assumed to be 160 litres waste
production/day per person. This is a post-septic treatment to further
purify the water and nutrients. Once the liquid reaches near the gravel
level, water and nutrient absorbing plants were planted. Assumed
hydraulic loading was 50 mm/day justified by higher temperatures in
Fiji. The operational concept of the wetland at Tagaqe was for blackwater
(including kitchen wastes) treatment via septic tank and bathroom
greywater (piping connected directly to wetland), to be treated via a
subsurface-flow gravel-bed. The septic tank was also fitted with Biotube
Effluent filters to improve efficiency. To facilitate sampling of final
effluent quality, proposed discharge for this demonstration system is to a
gravel-filled seepage channel. Operational systems could utilize open-
bottomed seepage zones in the final stage of the constructed wetlands.
71
Figure 20: Construction stages of the gravel bed wetland at Tagaqe Village
[Photos: courtesy of Chris Tanner]
(a) Briefing before beginning work (b) Original septic tank before work
(c) Addition of biotube filters to septic
tank
(d) Preparing piping for greywater
(e) Starting digging work on wetland (f) Additional assistance on digging
(g) Completion of wetland less plants (h) Fully grown sedge on wetland
72
At a depth of 0.4 m and porosity of 0.38, gives nominal resident time of
2.8 days. Kinetic modeling predicts Biological Oxygen Demand removal of
around 70 percent and total nitrogen removal of 40-50 percent at full
loading. However, performance may improve at a lower loading rate of
less than 15 daily occupants (Tanner, 2004: personal communication).
The total cost of the wetland at Tagaqe was around FJ$6,000.
Figure 21: Similar cross section diagram of Tagaqe Village wetland [adapted from EPA, 2000]
4.3. Best-practice commercial wastewater treatment systems
4.3.1. Packed Bed Filter technology
Packed bed filters incorporating naturally occurring treatment media
such as sand and gravel have been used successfully for treating small
to medium volume wastewater flows for decades. Over the past fifteen
years, two types of packed bed sand filters have been most commonly
used—the single-pass filter and the recirculating filter. Single-pass sand
filters are capable of treating septic tank effluent to advanced wastewater
73
treatment levels or better. Single-pass filters have been most successful
with low influent strength. With high influent strengths, maintenance
may increase, although with a diligent service and monitoring program,
performance is not expected to suffer. Recirculating (multiple-pass) filters
also treat septic tank effluent to advanced wastewater treatment levels or
better. Multiple-pass recirculating sand/gravel filters have been most
popular in applications with medium to large wastewater flows (Bounds,
2002).
The purpose of a sand filter is not only to remove sediment and
suspended solids, but mainly to provide biological aerobic treatment of
the wastewater. This is considered secondary treatment. The first six
inches of the filter is where most biological treatment occurs. Here is also
where suspended solids and BOD are removed. Sand filters are very
effective at lowering suspended solids as the media in the filter holds
onto the solids well. Faecal coliform bacteria removal ranges from 99-
99.99 percent (Orenco, 2009).
Recirculating filters, like single pass units pre-treat effluent from septic
tanks before it is released into the environment. Instead of all the water
in the underdrain flowing to the soil absorption system, the pipes return
some of it to the recirculation/dosing tank. Here the water is combined
with the effluent from the septic tank. Recirculating filters can be smaller
74
and have less odour than single pass units. Recirculating filters use a
more coarse material and have a higher hydraulic loading than single
pass units therefore reducing the removal of faecal coliform bacteria.
However, almost all ammonia is removed through nitrification (Orenco,
2009).
4.3.2. Limitations of Sand/gravel Media Packed Bed Filters
While sand/gravel media Packed Bed Filters are an excellent choice for
wastewater pretreatment, certain limitations have prevented them from
being considered at all sites:
Land area — some sites lack the land area required for a sand
filter. Single-pass sand filters for single-family homes typically
require between 300 and 400 square feet, depending on
jurisdictional design or flow criteria.
Media quality and accessibility — good quality sand media is
occasionally not locally available, resulting in either high
transportation costs or the use of inferior local media. In addition,
getting sand to some sites—such as islands, mountainous regions,
or other isolated areas—can be difficult.
Installation quality — sand filters are typically built onsite with
locally available materials, and the quality of installation is
partially contingent on the consistency of these materials, and the
knowledge and ability of the installing contractor.
75
Serviceability — the ease of maintaining a buried onsite single-
pass sand filter has been a long-term design concern that resulted
in robust designs with low loading rates. The low loading rates are
intended to ensure 10 to 20 years of continuous usage with little to
no intrusive filter maintenance because replacing the sand media
can be difficult and costly (Bounds, 2002).
4.3.3. Textile-based Packed Bed Filters
The efforts to improve loading capacities and serviceability have led to
extensive research into a wide variety of media (e.g., foam, glass, styrene,
plastic products, expanded clays, zeolite, limestone, furnace slag, peat,
etc.). Over the past decade, this research has led to the development of
an advanced technology for packed bed filters that uses an engineered
textile medium assembled in a variety of configurations. Textile provides
all the benefits inherent in the packed bed filter design but overcomes
the limitations listed above (Bounds, 2002).
Advantages of Textile-based Packed Bed Filters
Land area — the land area needed is significantly smaller than
that for sand filters because loading rates are 5 to 30 times higher.
Thus, the footprint area for a textile filter serving a typical four-
bedroom single-family home is now only about 20 square feet. If
the textile filter is positioned over the processing tank, virtually no
additional area is required.
76
Media quality and availability — the manufactured textile medium
ensures consistent quality and availability.
Installation quality — lightweight textile medium and small filter
size make pre-manufactured treatment units practical, eliminating
onsite construction and reducing installation time, labour, and
construction errors. These characteristics make textile systems
ideal for cost-saving self-help programs and particularly suited for
difficult-to-access and remotely located sites.
Serviceability — special configurations allow for ease of
maintenance and cleaning without expensive or large excavation
equipment, or the need for replacing the medium (Bounds, 2002).
Porosity — the porosity of the textile media is several times greater
than that of sand, gravel, and other particle-type mediums. The
more porous the medium, the greater its hydraulic conductivity,
the greater its air space (which enhances the capacity of passively
ventilated systems and free air movement), and the greater its
capacity for the accumulation of solids and biomass development.
Surface area — textile media can be blended with a variety of fibers
to achieve relatively large total surface area per unit volume.
Expanding the biomass growth area provides a greater surface
potential for air and effluent to interface and come in contact with
the biomass.
77
Water holding capacity — the water-holding capacity of textile
media also varies considerably depending on the media density,
type of material, and blend of fibers. The water-holding capacity in
textile media is also several times greater than expected in the
sands and gravels used in filters. Water-holding capacity performs
a key function in the treatment process. Together with the
programmed dosing time and frequency, it governs the effluent
retention time within the filter and ultimate effluent quality
(Bounds, 2002).
4.3.4. AdvanTex AX100 treatment systems
In packed bed systems, effluent trickles through a moist, porous media.
Microorganisms growing in the media remove impurities from the
effluent. Compared with advanced treatment systems that use liquid or
membranes, packed bed systems use less electricity, require less
maintenance, and are much less prone to upsets from abuse. For
advanced treatment, the AdvanTex Treatment System is ideal. AdvanTex
systems can reduce BOD, TSS, and faecal coliform by up to 98 percent
while nitrogen can be removed by up to 70 percent (Orenco, 2009).
AdvanTex systems possess the characteristics of an aeration treatment
unit. The filter itself is a textile material contained in a fibre-glass unit.
The components of the system include a septic tank, the AdvanTex unit
itself, and a drain field. Like sand filters, AdvanTex are biological reactors
78
which rely on bacteria for treatment of the wastewater. The textile of the
filter is where the colonies of bacteria grow and treat the water. Typical
effluent from AdvanTex systems contains BOD, suspended solids, and
total nitrogen levels of less than 10 mg/L. For proper operation AdvanTex
systems should be maintained annually (University of Wisconsin, 2009).
The AdvanTex AX100 onsite Treatment System is ideal for multi-family
residential properties; cluster community systems; subdivisions, resorts
and golf course developments; mobile and manufactured home
communities; parks and rest areas; truck stops, restaurants and
casinos; and school or office buildings (Orenco, 2007).
The patented AdvanTex Treatment System (Figure 22) is a recirculating
filter that has been configured like a recirculating sand filter (RSF).
Similar to recirculating sand filters, AdvanTex is affordable, reliable and
low maintenance. It is superior to other packed bed filters, however, in
its serviceability and longevity (Orenco, 2007).
In addition, the AdvanTex system is also superior in its treatment media.
AdvanTex uses a highly efficient, lightweight textile that has a large
surface area, lots of void space, and a high degree of water holding
capacity. Consequently, AdvanTex Treatment Systems can provide
treatment equivalent to that of sand filters at loading rates as high as
79
2,000 mm/day or one fortieth the footprint of the wetland. This implies
that AdvanTex can treat high volume commercial and multi-family flows
in a very compact space (Orenco, 2007).
Figure 22: the AdvanTex AX100 Treatment System at Crusoe’s Retreat
4.3.5. Description of Crusoe’s wastewater treatment system
The treatment system at Crusoe‘s Retreat is a simple, relatively low cost
and highly efficient wastewater treatment package set up in collaboration
between scientists at the National Institute of Water and Atmospheric
Research (NIWA), the University of the South Pacific and Innoflow
Technologies Limited, as an attempt to reduce effects of effluent disposal
on the coastal environment. As a result, the Crusoe‘s Retreat between
Navua and Sigatoka agreed to install this system for their small resort.
80
Installation of the AdvanTex AX100 wastewater management system
and construction of all tank work was completed in April 2005 by local
workers, under the supervision of a staff from Innoflow Technologies Ltd.
The system re-used existing septic tanks, after thorough investigation to
ensure suitability, with two lower tanks feeding to a new pumped tank
and an upper tank that flows directly via gravity into the recirculation
tank. All tanks were also fitted with Biotube Effluent filters to improve
efficiency (Figure 23). Although the AdvanTex system is usually installed
below ground, the site constraints at Crusoe‘s Retreat did not allow for
burying the treatment plant. As a result the system was just installed
above ground, as overall performance would not be affected (Innoflow,
2005).
Figure 23: Biotube Effluent filters used in septic tanks at Tagaqe and Crusoe’s
Retreat [Photo: courtesy of Chris Tanner]
The Crusoe‘s treatment system was estimated for 13 two-people ‗bures‘
(bungalows) with larger allowance for another additional four two-people
81
―bures‖, due to increased water consumption in resorts (longer showers)
at a peak flow of 7,000 litres/day. A pump station with a capacity of
7,000 litres was constructed close to the beach to collect wastewater
from lower septic tanks and transport up to the recirculation tank with a
size of 7,000 litres. Recirculating textile packed bed reactors (rtPBRs)
with a surface area of 12m2 were chosen to ensure anticipated
performance (see schematic diagram in Figure 24).
Figure 24: A schematic diagram of the Crusoe’s treatment system as
built
Upper septic
tank fitted with
biotube filters
Recirculation Tank
New lower
septic tank
with biotube
filters & pump
S h o r e l i n e
Flower beds
Crusoe’s Entrance,
Reception & Car Park
Flow through gravity
Flow via pump
Carbon filter &
ventilation fan
AX
10
0 T
extile F
ilter PO
D
Gravity discharge
Point to flower beds
Old septic tank
Old septic tank
82
The operational concept was that blackwater and greywater would first
be treated initially via three septic tanks. Wastewater from one upper
septic tank would then flow into the recirculation tank due to gravity,
whilst wastewater from two lower septic tanks would be electrically
pumped up into the recirculation tank. The effluent pumping system at
the recirculation tank would then pump the wastewater into the Textile
Filter Pod and then comes back into the recirculation tank through an
underlain collection pipe. The recirculation pumps were anticipated to
run for 2.1 hours per day at 0.75 kW per pump at peak periods. The
process from the recirculation tank through to the Textile Filter Pod and
then back to the recirculation tank takes 4 cycles, before the treated
water is eventually discharged into a flower garden by gravity (Figures
26-30). It was anticipated that a reasonably significant outflow up to
about 6,500 litres per day of treated wastewater would be discharged on
the ground disposal system and flower garden (Innoflow, 2005).
According to an Innoflow engineer, Chris Shortt, who supervised the
instalment stage the system basically consisted of three major phases
(cited in Hasan, 2005).
a) Primary treatment – takes place inside septic tanks (already
existing ones and a new one built). These were modified slightly by
installing a filter at the outlet so that big solid matter (>3mm) do
not pass through and move to the next phase.
83
b) Secondary treatment – the wastewater from the septic tank gets
filtered again in a small chamber and moves to a large tank for
treatment. In this tank are specially designed filters which are
used instead of the traditional sand filters. Its advantage is that it
is more efficient for microbial growth and it will last a life time if
properly maintained. The wastewater is recirculated four times
through the filters before they move to the final phase.
c) Disposal – the treated waste is disposed to gardens for nutrient
uptake by plants.
The main nitrogen-related processes in the wastewater treatment system
are nitrification and denitrification. Basically, the ammonia in waste is
converted to nitrate (nitrification) and then the nitrate is changed to
nitrogen gas (denitrification). When wastewater is not properly treated, a
lot of nitrate is discharged into the sea and other water bodies like rivers
and streams leading to algae problems. The treatment system set up at
Crusoe‘s was anticipated to transform about 75 percent of nitrate to
nitrogen gas and about 15 percent was assumed as going to be taken up
by the plants. Hence only 10 percent would be going out to the sea which
will help keep the reef ecosystem healthy (Hasan, 2005). The rate of
nutrient removal was expected to exceed 90 percent while the treatment
performance for suspended solids and biological oxygen demand were
predicted to be less than 15 mg/litre (Innoflow, 2005). Total cost for the
84
Crusoe‘s wastewater treatment system was about FJ$62,500 which was
cheaper than other systems used by nearby hotels.
Figure 25: (a) the ProSTEP Effluent pump switchboard at the recirculation tank; (b) the lower septic pumping system closer to the beach at Crusoe’s
Figure 26: (a) George Reece standing beside the Carbon filter and ventilation fan
of the treatment system at Crusoe’s; (b) The AX100 textile filter pod fibre layers
for wastewater treatment at Crusoe’s Retreat
(a) (b)
(a) (b)
85
Figure 27: (a) The recirculation splitter valve and the effluent pumping system; (b)
the recirculation splitter valve with piping connections from septic tank effluents
and the AX100 treatment system
Figure 28: The flower gardens and ground disposal area at Crusoe’s Retreat
4.4. Greywater treatment “drum system” experiment
4.4.1. Experimental Plan
The model drum experiment was aimed at evaluating the performance of
the greywater treatment system under two different hydraulic loading
rates equivalent to one or two drums per household.
Rather than the daily hydraulic load being distributed evenly throughout
a day, greywater from a household is typically generated in episodic
pulses (e.g. when washing machine discharges or someone has a
(a) (b)
86
shower). It is expected that the treatment efficiency of a greywater system
(and underlying soil) will be different for large pulses of inflow with rest
periods in between compared to small continuously trickling inflows.
Thus, a secondary aim of the experiment was to assess the performance
of the systems under different dosing regimes (that is, by dividing the
―average‖ daily hydraulic load received by each mesocosm into doses of
different size and frequency). It was proposed that two greywater
treatment mesocosms be set-up and operated side-by-side at two
different loading rates as depicted in Figures 29 - 31. Specifically, High
Loading Rate is equivalent to one drum system per ―typical‖ household.
Low Loading Rate is equivalent to two drums per household.
The mesocosms were set-up and operated for at least 4 weeks before the
start of monitoring in order to allow the build-up of micro-organisms,
bio-films and potentially clogging organic matter within the systems.
4.4.2. Hydraulic Loading Rates
Estimated daily greywater load from Votua household:
Total wastewater load from Votua household ≈ 250 L/person/day
multiply by 5 people/household = 1250 L/day
Assuming greywater load = 70% of total wastewater hydraulic load, then:
Greywater hydraulic load (Q) = 875 L/household/day
= 0.875 m3 day-1
87
The areal hydraulic loading rate (HLR) for one greywater drum per
household (High Loading Rate) = [Q] / [surface area of drum (≈ 0.24 m2)]
= 0.875 m3 d-1 / 0.24 m2
= 3.65 m day-1
= 3650 mm/day
= the ―High Loading Rate‖
The areal HLR for two greywater drums per household (Low Loading
Rate): = 0.875 m3 d-1 / 0.48 m2
= 1.82 m day-1
= 1820 mm/day
= the ―Low Loading Rate‖
Thus, if the surface area of the 20L buckets used in the mesocosms is
0.07 m2 (assuming diameter of bucket = 30 cm), the daily hydraulic load
(amount of greywater to be added each day) for the two mesocosms will
be: High Loading Rate = 3.65 m day-1 * 0.07 m2
= 0.255 m3 day-1
= 255 L/day
Low Loading Rate = 1.82 m day-1 * 0.07 m2
= 0.127 m3 day-1
= 127 L/day
During the 4 week ―start-up‖ period, the greywater mesocosms were
dosed with artificial greywater twice per day (morning and afternoon).
88
Thus, at each dose the High Loaded mesocosm received 127.5 L and the
Low Loaded mesocosms received 63.5 L of greywater.
Figure 29: (a) an in situ greywater treatment drum system at Votua Village along the Coral Coast; and (b) an ex-situ model at the university
(a) (b)
89
Figure 30: Side view of greywater treatment mesocosm. All lengths in cm
[diagram: courtesy of Tom Headley, 2007]
Ventilation pipe: 20mm PVC, ends open, sitting on top of gravel under-drain.
200L plastic drum
2 x 20L plastic buckets (usually white), 35cm high by 30cm diameter: One sitting partially inside the other, each one filled with 25cm depth of media. Top bucket with 2cm holes drilled in bottom to allow drainage through to Bottom bucket which has 1cm holes drilled in a band around the sides (no holes in bottom).
30cm depth of coconut husk
pieces (approx. 10cm x 5cm)
30cm depth of coconut shell pieces.
Break shell into quarters with hammer
Coral rock: 5-10cm diameter rocks
Sandy soil: Soil typical of Votua, placed in drum to resemble in-situ soil.
Gravel under-drain: layer of 10mm gravel to allow free drainage.
50mm PVC drain pipe: plumbed through the bottom of drum (at lowest point so that water does not sit in bottom of drum) using a “tank connector”. Set-up so that effluent goes to sewer, but also to enable sample collection.
Greywater inflow
Bricks or concrete blocks: to enable outlet to be placed in bottom of drum.
40
40
90
10
10
10-15
90
Figure 31: Details of drainage holes in 20L buckets used to contain coconut shell
and husk [diagram: courtesy of Tom Headley]
4.4.3. Targeted greywater characteristics
Greywater is the wastewater generated from showers, bathtubs, hand
basins, laundry, washing machines and kitchen sinks and consequently
contains a mixture of soaps, detergents, food particles, fats, oils, soil,
hair, and potentially some small amounts of faecal matter and urine.
Studies (Urban Water Research Association of Australia, 1996; Jefferson
et al., 1999 & 2001; Eriksson et al., 2002; Brown and Palmer, 2002; Toifl
et al., 2006) show that greywater has a similar organic strength to
domestic wastewater, but relatively low suspended solids (i.e. greater
proportions of the contaminants are dissolved).
1cm diameter holes drilled at approx. 5 cm spacing around bottom bucket in a 25cm high band from upper level of coconut shell down to 5cm from bottom of bucket
30 cm
5 cm 0 cm
Upper bucket with 2cm holes drilled in bottom. Bucket sits on top of coconut shell in bottom bucket.
91
Table 5 provides a summary of typical greywater characteristics from
other studies, which was used as target concentrations for the artificial
greywater (Toifl et al., 2006).
Table 5: Summary of typical greywater characteristics targeted in the experiment
Parameter Target greywater concentration
Suspended solids 60 – 80 mg/L BOD 150 – 200 mg/L
Temperature 20° - 30°C pH 6.5 – 8.0 Turbidity 60 – 80 NTU Sodium 80 – 130 mg/L Total Phosphorus 10 – 20 mg/L Total Kjeldahl nitrogen 3.0 – 5.0 mg/L Conductivity 450 – 550 μS/cm COD 250 – 400 mg/L TOC 100 – 150 mg/L Total coliform 103 – 104 counts/100ml
E.coli 101 – 102 counts/100ml
92
Chapter 5 Methodology
5.1. Introduction
This chapter outlines the field sampling procedures of wastewater,
freshwater and coastal seawater samples both from the wastewater
treatment initiatives, freshwater creeks, and near shore sites along the
Coral Coast of Fiji. It also attempts to detail the sampling plan and recipe
for the greywater drum system experiment. Analytical techniques for
varying water quality parameters are also explained.
5.2. Field sampling procedures
5.2.1. Dissolved inorganic nutrients
The frequency of sampling was done on a monthly basis in average. For
the constructed wetland at Tagaqe, there were eight samplings
undertaken on 15th June 2005, 18th July 2005, 20th October 2005, 11th
May 2006, 8th June 2006, 5th July 2006, 14th September 2006, and 8th
October 2006. In regard to the Crusoe‘s wastewater treatment system,
five sampling trips were completed on 20th October 2005, 8th June 2006,
5th July 2006, 13th August 2006, and 14th September 2006. For general
Coral Coast sites, five sampling trips were carried out on 18th June 2005,
20th October 2005, 11th May 2006, 8th June 2006, and 5th July 2006.
Votua Creek sampling was done on 8th June 2006, 5th July 2006, 13th
August 2006 and 14th September 2006 (refer to Appendix).
Unfortunately, the loading rates or flow rates for each device was not
determined due to lack of funding.
93
Water samples for dissolved inorganic nutrients including ammonia
(NH3), nitrate (NO3-), nitrite (NO2
-) and phosphate (PO43-) were filtered in
the field and collected in acid-cleaned polypropylene bottles with leak-
proof caps. Prior to collection, each sampling bottle and cap was soaked
overnight in 10 percent hydrochloric acid bath and later washed with
deionised water in the laboratory to remove bacteria. While in the field,
each sampling bottle was rinsed at least three times with sample solution
before the final sample was collected. For the Tagaqe wetland and the
wastewater treatment system at Crusoe‘s Retreat, filtered wastewater
samples were collected at the inlet and outlet pipes from both treatment
initiatives (Figures 32-34). In general nearshore seawater and freshwater
creek sampling sites, filtered samples were collected at a depth of about
10-50 cm below the water surface. Filtration of wastewater and seawater
samples for determination of dissolved nutrients was carried out using
the Whatman GF/C, 1.2 µm pore size filters to remove large particles,
plankton and bacteria. This was conducted in-situ using the suction
filtration kits. During transportation to the laboratory samples were kept
in ice and upon arrival they were either analysed immediately (i.e.
ammonia) or stored by freezing the filtrate at less than 4 ºC for up to 7
days in order to attain good results.
94
Figure 32: Acid bath for field sampling bottles and reagent/standard preparation
Figure 33: Field sampling at Crusoe’s wastewater treatment system
Figure 34: (a) Sample collection at the Tagaqe wetland inlet; (b) sample collection
at the Tagaqe wetland outlet
(a) (b)
95
5.2.2. Bacteria
The specific bacteria that were monitored in water samples from this
study include faecal coliform and E/coli. For coliform counts, the
wastewater and/or seawater samples were collected in either a labeled
500ml or 250ml sterilised sampling bottle. Wastewater samples were
usually collected in 250ml sterilised bottles whilst 500ml bottles were
used for freshwater or seawater samples. Collected samples were placed
in ice and transported back to the micro laboratory for immediate
analysis in order to achieve accurate results.
5.2.3. In-situ measurements
Several water quality parameters including salinity, temperature,
conductivity, and dissolved oxygen (DO) were measured in-situ using a
calibrated measuring probe attached by a 5m cable to a conductivity
meter (model: YSI-85).
5.2.4. Other water quality parameters
Water samples intended for analyses including total suspended solids
(TSS), Total Phosphorus (TP), Total Nitrogen (TN), Total Kjeldahl Nitrogen
(TKN), Biological Oxygen Demand (BOD), and pH measurement were
collected in unfiltered labeled acid-cleaned bottles and kept in ice whilst
transported back to the laboratory. BOD and pH analyses were
undertaken immediately whilst samples for other parameters could be
refrigerated for up to 7 days.
96
5.3. Greywater treatment drum system experiment
5.3.1. Water Quality Monitoring Program
After the completion of the drum system experiment set up in mid April
2007, as explained in Chapter 4, a start up period of at least four weeks
was initiated whereby monitoring was conducted on the two mesocosms
under various dosing regimes (Table 6).
Table 6: Preliminary water quality monitoring program to compare different
loading regimes [courtesy of Tom Headley]
Monitoring Mesocosm Average Daily HLR
Dosing Regime#
Dose volume
Period at this regime before sampling
Period Loading Rate
L/day doses/day L/dose (minimum)
1 (large High 255 3 85 3 days
doses) Low 127 3 42.5
2 (moderate
High 255 6 42.5 2 days
doses) Low 127 6 21.25
3 (small High 255 12 21.25 1 day
doses) Low 127 12 10.6
# assumes mesocosms will be dosed manually during work hours
[between approx. 8am and 5pm, (9 hour period)].
5.3.2. Other Monitoring
Hydraulic performance:
The most likely mode of long-term failure of this type of greywater
management system was clogging or blockage of the soil surface where
the partially treated greywater infiltrates into the natural soil. A ―bio-
mat‖ of biofilm, slime and accumulated organic solids was attributed to
be the main cause of such clogging. The degree of clogging was expected
to be greatest in the highly loaded mesocosm. A gauge of the degree of
soil-interface clogging was obtained by comparing the hydrograph (i.e.
97
effluent flow rate versus time) for the two mesocosms following
application of a single dose, both when first set-up (i.e. clean system) and
at the end of the experiment. It was anticipated that the time taken for a
dose to drain through the system to increase as the soil-interface
becomes clogged.
Visual Observation of the condition of the different layers:
At the very end of the experiments, the various components of the system
were visually inspected for signs of clogging, organic matter build-up and
decomposition.
5.3.3. Starting recipe for artificial greywater solution
Table 7 provides a starting recipe for the artificial greywater solution
used. This starting recipe can be refined after preliminary analysis of the
drum system during the start up phase. Where necessary, urea could
also be added to increase the nitrogen concentration.
Table 7: Starting recipe for artificial greywater [courtesy of Tom Headley]
Ingredient Amount (per litre) Relevant inputs
Soap (powdered or grated) 50 mg/L Na, BOD, COD, fat
Shampoo/dishwashing liquid 0.5 ml/L Na, BOD, COD, surfactant Secondary treated sewage 1.0 ml/L Bacteria, BOD Whole milk powder 200 mg/L BOD, COD, N, oils, fats Laundry powder 250 mg/L Na, B, P, surfactants Soil 25 mg/L TSS, microbes
5.4. Analytical methods
The analyses of dissolved chemical species including nitrate, phosphate,
ammonia, and nitrite in wastewater, seawater and/or freshwater samples
were performed using an automated Lachat Quik Chem Flow Injection
98
Analyser (FIA) based at the University‘s Institute of Applied Science
laboratory (Figure 35). The flow injection analyser uses colorimetric
analyses and has the ability to process up to 60 samples an hour with
high reproducibility using relatively small sample sizes of ~3 ml per
analysis (Loder, 2000). Nitrate and/or nitrite analysis was adapted from
the methods written by Diamond (1994), as detailed in Lachat
Instrument Quik Chem Method 31-107-04-1-A. Phosphate analysis
followed Lachat Instrument Quik Chem Method 10-115-01-1-B (Prokopy,
1994). Ammonia analysis was based on the methods written by Ninglan
(2002) and detailed in Lachat Instrument Quik Chem Method 31-107-06-
1-B. The accuracy and precision of each of the methods is shown in
Table 8 and their detection limits in Table 9. Data obtained for the
General Coral Coast water quality samples were statistically analysed
using a software package known as SPSS 16.0 for Windows.
Table 8: Accuracy and precision for each Lachat Quik Chem FIA method
Nutrient species
Method (reference) Conc (µmol/L)
Mean conc (µmol/L)
Standard deviation
% RSD
Nitrate/
nitrite
31-107-04-1-A
(Lachat Inc., 1994)
10.0 10.1 0.090 0.89
Ammonia 31-107-06-1-B (Lachat Inc., 2002)
1.0 1.0 0.04 3.5
Phosphate 10-115-01-1-B (Lachat Inc., 1994)
1.0 0.98 0.04 4.0
Nitrite 31-107-04-1-A (Lachat Inc., 1994)
0.50 0.49 0.011 2.24
99
Table 9: Method detection limit for each Lachat Quik Chem FIA method
Nutrient species
Method (reference)
Known conc. (µmol/L)
Mean conc. (µmol/L)
Standard deviation
Method Detection Limit (µmol/L)
Nitrate/nitrite 31-107-04-1-A (1994)
0.28 0.25 0.015 0.04
Ammonia 31-107-06-1-B (2002)
3.57 3.64 0.138 0.27
Phosphate 10-115-01-1-B (1994)
0.50 0.48 0.005 0.01
Nitrite 31-107-04-1-A (1994)
0.25 0.24 0.005 0.01
Typically, four standards (high, medium, low, and zero) were used in the
colorimetric analysis to create a standard curve from which the unknown
sample concentration was determined. These standard curves generally
had an r-value of 0.9998 or higher in order to be accepted. Both a blank
and known standard were sampled at least every ten samples throughout
each analysis to monitor instrument drift and base level consistency.
Later, during the data processing, these blank sample concentrations
were averaged together and subtracted from the standard and unknown
samples. The standard concentration values were compared to confirm
overall analytical accuracy as well as stability across the run.
Other water quality parameters were analysed using internationally
accredited ‗Standard Operating Procedures‘ at the Institute of Applied
Science laboratory, as detailed in the American Public Health
Association‘s (APHA) Standard Methods for the Examination of Water
and Wastewater (Clesceri et al., 1998; APHA, 2005). For instance, faecal
100
coliform (method reference number APHA 9221-B); E/coli (APHA 9222-G);
total suspended solids (APHA 2540-D); biological oxygen demand (APHA
5210-B); pH (APHA 4500-H); total Kjeldahl nitrogen (APHA 4500-Norg);
total phosphorus (Danish Std); and total nitrogen (APHA 4500-N).
Figure 35: The Auto Sampler Injector which sucks sample to be passed through
the FIA manifold
5.5. Automated Flow Injection Analysis
5.5.1. General description
Flow Injection Analysis (FIA) is a continuous flow technique which is
ideally suited to rapid automated analysis of liquid samples. In a flow
injection analyser, a small, fixed volume of a liquid sample is injected as
a discrete zone using an injection device into a liquid carrier which flows
through a narrow bore tube or conduit. The sample zone is progressively
dispersed into the carrier, initially by convection, and later by axial and
radial diffusion, as it is transported along the conduit under laminar flow
conditions. Reagents may be added at various confluence points and
101
these mix with the sample zone under the influence of radial dispersion,
to produce reactive or detectable species which can be sensed by any one
of a variety of flow-through detection devices (Figure 36). The height or
area of the peak-shaped signal thus obtained can be used to quantify the
analyte after comparison with the peaks obtained for solutions
containing known concentrations of the analyte (McKelvie, 1999; Lachat,
2003).
The complete process of sample/standard injection, transport, reagent
addition, reaction and detection can be accomplished very rapidly
(seconds to tens of seconds), using minimum amounts of sample and
reagents, and with excellent reproducibility (e.g. coefficient of variation,
CV, generally < 2 percent). Although complete equilibrium may not be
achieved during this process, quantitation is possible because both
standards and samples are dispersed to the same extent and processed
in an identical manner (McKelvie, 1999).
102
Figure 36: (a) The injector and sample zone; (b) reagents being added to samples;
(c) analyser pumps; (d) a 4 channel manifold; (e) nitrate column on manifold; (f)
the computer system that log results; (g) FIA waste outlet; (h) peak shaped signals on computer for analyte
(a) (b)
(c) (d)
(a)
(e) (f)
(g) (h)
103
5.5.2. Components of a Flow Injection Analyser
The assemblage of flow tubing, mixing coils, injection valve, etc, in a
given configuration, used in a flow injection analysis system is referred to
collectively as a manifold. A typical flow injection analysis manifold is
shown in Figure 37 using the common symbolic notation (McKelvie,
1999; Lachat, 2003).
Figure 37: Schematic diagram of a typical flow injection analysis manifold. P is a
pump, C and R are carrier and reagent lines respectively, S is sample injection,
MC's are mixing coils, D is a flow through detector, and W is the waste line.
The major components of a flow injection analysis system (Ruzicka and
Hansen, 1975; Valcarcel and Luque de Castro, 1987; Ruzicka and
Hansen, 1988; Karlberg and Pacey, 1989; Burguera, 1989; McKelvie,
1999) are:
1. A ‗propulsion‘ system for delivery of carrier and reagent solutions.
This is most frequently a peristaltic pump or pumps, typically with
the capacity to pump between four to eight carrier/reagent lines.
104
2. ‗Flow tubing or conduit‘ with an internal diameter of 0.3 -1.0 mm
is generally used to maintain laminar flow conditions and
controlled dispersion of the sample zone.
3. A ‗sample injection‘ device for reproducibly introducing a small
volume of sample into the flowing carrier or reagent stream.
4. A ‗mixing device‘ to promote radial diffusion, and hence reaction
between sample zone and reagents. Mixing devices usually consist
of coils of tubing. If a small coil radius is used, secondary flow
occurs and enhances the radial mixing with minimal gains in axial
dispersion.
5. One or more flow-through ‗detection devices‘ which may sense
changes in absorbance, fluorescence, chemiluminescence, atomic
emission or absorption, infra-red absorption, pH, electrode
potential, diffusion current, electrical conductivity, turbidity, mass,
etc.
6. An ‗instrument control-data acquisition/processing/output
system‘. An automated FIA system may be controlled using a PC
with a suitable analogue to digital board, and control/data
acquisition software.
5.5.3. General guidelines for using the FIA for nutrient analysis
The following guidelines were developed by the researcher during the
study to ease future usage of the Flow Injection Analyser at the
105
University of the South Pacific‘s Institute of Applied Science laboratory in
regard to nutrient analytes in water samples:
Manifold Connections: Nitrate cannot be interchanged. Only
interchange ammonia and phosphate manifold and tubing (refer to
Quik Chem Methods).
Heating temperatures: phosphate - 37˚C; nitrate/nitrite - 33˚C;
and ammonia - 60˚C.
To adjust temperature: Press menu button until ―SP‖ appears, use
up key (۸) or down key (۷) to adjust to appropriate temperature,
press ENTER to save, press menu button once to show the current
temperature, run machine with Deionised water for 15 minutes to
reach appropriate temp.
Ammonia, nitrate/nitrite and phosphate could be analysed in one
run. Always analyse for ammonia when sample is still fresh whilst
nitrate/nitrite and phosphate could be analysed later after freezing
within 7 days.
For ammonia analysis, all reagent containers should be covered
with parafilm after insertion of the transmission lines to prevent
contamination from airborne ammonia. Also add standards and
samples slowly in sampler to avoid contamination from ammonia
in air. Ensure to add reagents in the order that they appear on the
Lachat Quik Chem Method 31-107-06-1-B manifold to reduce
106
staining. Also prepare fresh reagents and ensure that EDTA in the
buffer is completely dissolved.
5.5.4. General steps for running samples on the FIA
1. TURN ON: main power-point, pump switches, manifold switch and
computer (data system).
2. Pump DI water through all reagent lines for at least 15 minutes to
reach appropriate temperature and to check for leaks and smooth
flow.
3. Switch to reagents and allow 5 minutes at ―Manual rate pumping‖
for system to equilibrate for nitrate/nitrite and phosphates. Keep
the nitrate channel column switch at ―Off line‖ to avoid bubbles
from entering the Cadmium Column at this stage. For ammonia:
allow 15 minutes for system to equilibrate.
4. Place 4-5 tubes of blanks (Deionised water) in the sampler, open
Omnion 3.0 program in computer or any existing file e.g: ‗Exsley‘ in
either ‗My computer‘ or ‗My documents‘. Type Deionised water in
the spaces provided beside each cup number and click ‗START‘.
Now turn the Cadmium column switch to ―On line‖ and system will
start running after 90 seconds. A low stable baseline needs to be
achieved and should not give a peak. If the blank peak is negative,
the carrier is contaminated. If the blank peak is positive, the blank
is contaminated.
107
5. Place respective nutrient standards and/or samples in the
sampler. Input the information required by the data system such
as concentration, cup number, replicates, check standards,
blanks, moos, field blanks and so forth. Click Start and injection of
standards and/or samples will start within 90 seconds. In certain
cases it might go beyond the 90 seconds.
6. Verify calibration using a midrange calibration standard every 10
samples or every analytical batch. Use 35ug/L for nitrate, 10ug/L
for phosphate, and 20ug/L for ammonia.
7. At the end of analysis, click STOP and select appropriate mode of
stopping. Turn pump to ‗minimum setting‘ when the system shows
―IDLE‖ in the computer and check for calibration curves. Discard
way off points to get good calibration curves.
8. Click on icon for graph enlargement on the bottom left hand corner
of the computer and left click on graph and move bars to detect
peaks. Right click on graph and select ―re-run peak detection
limits‖.
9. Flash all tubes with DI water for at least 10 minutes while turning
the nitrate channel column to ―Off line‖. Note: For phosphate
(optional); place the colour reagent and ascorbic acid transmission
lines into NaOH-EDTA solution and pump this solution for 5
minutes while placing the other lines for nitrate/nitrite and
108
ammonia separately in deionised water. Then place all lines in
deionised water and pump for another 5 minutes.
10. Save Report or Print customized copy on printer installed and close
computer and turn off all switches including the main power-point.
Ensure to release all the pump clippings to relieve unnecessary
stress on the tubing.
109
Chapter 6 Results
6.1. Introduction
This chapter progressively presents the results and monitoring data from
the Tagaqe Village constructed wetland, Crusoe‘s Resort treatment plant,
general Coral Coast sites, Votua Village Creek monitoring, and an ex-situ
model drum experiment.
6.2. Tagaqe Village constructed wetland
The subsurface constructed wetland at Tagaqe Village was completed in
December 2004 and left for about six months to allow for the wetland
plants to flourish and the system to stabilise. Preliminary monitoring
occurred in May 2005, but actual monitoring commenced in June 2005
and ended in October 2006 (Appendix A). The frequency of sampling was
done on a monthly basis in average. There were eight samplings
undertaken on 15th June 2005, 18th July 2005, 20th October 2005, 11th
May 2006, 8th June 2006, 5th July 2006, 14th September 2006, and 8th
October 2006. The overall wetland monitoring period was around sixteen
months and a summary of the water quality results are presented in
Table 10. A visual observation of untreated and treated wastewater from
the wetland is shown in Figure 38.
110
Figure 38: Sample of treated and untreated wastewater from Tagaqe wetland
Table 10: Summary of mean water quality results from Tagaqe wetland over the
period between June 2005 and October 2006 (n=8)
Parameters Influent Effluent % Removal
Temperature (˚C) 28.6 27.2 - Salinity (ppt) 0.0 0.0 - Dissolved Oxygen (mg/l) 0.16 0.82 - Conductivity (mS/cm) 0.12 0.02 - pH 6.80 6.85 - Faecal Coliform (c/100ml) 3.6 x 106 2.4 x 104 99.3 E.coli (c/100ml) 1.4 x 106 1.5 x 104 98.9 Total Suspended Solids (mg/l) 886.6 30.6 96.5 BOD5 (mg/l) 324.8 17.3 94.7 NH3-N (µmol/L) 4596.8 798.1 82.6 NO3-N (µmol/L) 1.04 6.08 - NO2-N (µmol/L) 0.14 0.07 50.0 Total Inorganic Nitrogen (µmol/L) 4598 804.3 82.5 Total Kjeldahl Nitrogen (µmol/L) 8169.2 1997.6 75.5 PO4-P (µmol/L) 392.0 96.1 75.5 Total Phosphorus (µmol/L) 539.0 166.6 69.1
TIN: P ratio 9 5
Results for the Tagaqe wetland showed that removal efficiency of faecal
coliform, E/coli, total suspended solids (TSS), and biological oxygen
demand (BOD) exceeded 90 percent. Nitrogen removal from the wetland
ranged from 50.0 percent for nitrite (NO2-N) to 82.6 percent for ammonia
(NH3-N) while nitrate (NO3-N) appeared to show no direct removal. In
regard to phosphorus, the removal rate was 75.5 percent for phosphate
Treated Sample - clear and
earthy smell
Untreated Sample - cloudy
and foul smell
111
(PO4-P) and 69.1 percent for total phosphorus (TP). The total inorganic
nitrogen (TIN) to phosphorus (PO4-P) ratio was 9 for the untreated
influent and 5 for the treated effluent.
6.3. Crusoe’s Resort wastewater treatment plant
The wastewater treatment plant at the Crusoe‘s Resort was completed in
April 2005. Monitoring spanned for around 11 months starting in
October 2005 until September 2006 (Appendix B). Results for the
treatment plant performance are summarized in Table 11.
Table 11: Summary of mean water quality results from the Crusoe’s wastewater
treatment plant over the period between October 2005 and September 2006 (n=5)
Parameters Influent Effluent % Removal
Temperature (˚C) 30.3 30.7 - Salinity (ppt) 0.0 0.0 - Dissolved Oxygen (mg/l) 0.17 4.20 - Conductivity (mS/cm) 0.74 0.02 - pH 6.76 6.90 - Faecal Coliform (c/100ml) 8.9 x 105 5.8 x 104 93.5 E.coli (c/100ml) 1.0 x 106 5.3 x 104 94.7 Total Suspended Solids (mg/l) 35.2 12.8 63.6 BOD5 (mg/l) 63.6 18.0 71.7 NH3-N (µmol/L) 2702.4 872.7 72.7 NO3-N (µmol/L) 19.3 6.07 68.5 NO2-N (µmol/L) 0.57 0.28 50.9 Total Inorganic Nitrogen (µmol/L) 3214.2 879.1 72.7 Total Kjeldahl Nitrogen (µmol/L) 3194.3 1347.6 50.1
PO4-P (µmol/L) 220.4 64.6 70.7 Total Phosphorus (µmol/L) 263.1 103.9 60.5
TIN: P ratio 12 8
Data from the Crusoe‘s wastewater treatment system outlined in Table
11 showed that the percentage removal of faecal coliform was 93.5;
E/coli 94.7 percent; total suspended solids 63.6 percent whilst 71.7
percent was attained for biological oxygen demand. Ammonia (NH3-N)
was removed at 72.7 percent; nitrate (NO3-N) 68.5 percent; and nitrite
112
(NO2-N) 50.9 percent. In addition, phosphate (PO4-P) removal reached
70.7 percent while total phosphorus yielded 60.5 percent. The nitrogen
to phosphorus ratio was 12 for influent and 8 for treated effluent.
6.4. Coral Coast nearshore water quality monitoring
Coral Coast nearshore sites consist of monitored locations adjacent to
tourist resorts and villages similar to sites studied by Mosley and
Aalbersberg (2003). Monitoring under this study was undertaken
between July 2005 and July 2006 (Appendix C). The average results are
detailed in Table 12. A comparative baseline data for the Coral Coast
which was obtained by the University of the South Pacific‘s Institute of
Applied Science researchers (e.g. Bale Tamata, Luke Mosley, Bill
Aalbersberg and Sarabjeet Singh) is displayed in Table 13.
Table 12 showed that salinity for the 17 sites monitored varied between
27.3 ppt at Korotogo Bridge and 34.1 ppt at Malevu Village shoreline.
Average salinity for the 17 sites was 32.0 ± 0.4 ppt. Water temperature
ranged from 27.2 ºC for Tabua Sands and Komave to 32.1 ºC for Naviti
Resort shoreline with a mean of 29.4 ± 0.3 ºC. Dissovled oxygen (DO)
levels were relatively moderate with a minimum of 3.47 mg/L at Korotogo
Bridge and a maximum of 7.13 mg/L at Outrigger Resort shoreline. The
mean DO level for the 17 sites was 6.10 ± 0.20 mg/L. Conductivity
fluctuated within 36.76 mS/cm (i.e. Matai Kadavu Beach) and 56.0
mS/cm for Malevu shoreline with an average of 49.92 ± 1.16 mS/cm.
113
In reference to nutrients, Komave and Tagaqe shorelines yielded the
lowest nitrate (NO3-N) values of 2.33µmol/L and 2.58µmol/L respectively.
The highest nitrate concentrations were observed for Malevu, Outrigger
Resort and Tubakula Resort shorelines with levels of 6.88µmol/L;
6.68µmol/L; and 6.18µmol/L respectively. However the average nitrate
level for the 17 monitored sites was 4.16 ± 0.35 µmol/L. Ammonia (NH3-
N) varied between 0.98µmol/L at Tabua Sands and 5.82µmol/L at
Korotogo Bridge with a mean of 2.09 ± 0.29 µmol/L. Site number 17 (e.g.
between Malevu & Vatukarasa) had the lowest nitrite (NO2-N)
concentration of 0.24µmol/L and the highest was observed at site
number 16 (Matai Kadavu Beach) with 0.63µmol/L. The average nitrite
value for the Coral Coast was 0.35 ± 0.02 µmol/L.
Phosphate (PO4-P) concentration was lowest at Hideaway Resort
(0.28µmol/L) and Fijian Resort (0.31µmol/L) shorelines whilst Votua,
Vatukarasa and Korotogo Bridge yielded the highest phosphate levels of
0.51; 0.58; and 1.14 µmol/L, respectively. A mean PO4-P of 0.43 ± 0.05
µmol/L was obtained for the Coral Coast water quality monitoring. The
total inorganic nitrogen (TIN) to phosphorus (PO4-P) ratio ranged between
9 and 29 with an average of 17 ± 1.
114
Table 12: Summary of mean water quality results from the Coral Coast between July 2005 and July 2006 (n = 5)
Site #
Place & (GPS Location) Sal (ppt)
Temp (ºC)
DO (mg/l)
Cond (mS)
NO3-N µmol/L
NH3-N µmol/L
NO2-N µmol/L
PO4-P µmol/L
TIN:P ratio
1 Fijian Resort (18-08.62S;177-25.76E) 33.0 29.2 5.77 53.45 3.09 1.71 0.38 0.31 17 2 Outrigger Resort (18-10.82S;177-
33.08E) 32.2 28.3 7.13 52.9 6.68 1.61 0.39 0.36 24
3 Tubakula Resort (18-10.86S;177-
33.46E) 32.7 29.1 7.06 51.95 6.18 1.25 0.37 0.33 24
5 Votua Village (18-12.69S;177-42.89E) 31.6 29.3 5.90 54.38 2.85 2.34 0.32 0.51 11 6 Tagaqe Village (18-11.91S;177-39.75E) 32.6 28.2 6.32 45.91 2.58 1.07 0.28 0.45 9 7 Sovi Bay (18-12.30S;177-36.39E) 33.4 30.1 5.86 53.97 3.74 2.14 0.33 0.35 18 8 Hideaway Resort (18-11.92S;177-
39.32E) 33.0 30.2 6.33 47.15 3.99 3.75 0.35 0.28 29
9 Naviti Resort (18-12.31S;177-41.84E) 31.5 32.1 5.97 51.83 3.59 3.07 0.29 0.37 19 10 Komave Village (18-13.38S;177-45.71E) 31.6 27.2 5.49 45.14 2.33 2.0 0.29 0.35 13 11 Tabua Sands (18-11.62S;177-37.89E) 32.1 27.2 5.95 51.63 3.90 0.98 0.30 0.33 16 12 Vatukarasa (18-10.85S;177-36.22E) 31.7 29.7 6.62 50.7 3.14 2.44 0.34 0.58 10 13 Malevu Village (18-10.85S;177-33.62E) 34.1 29.3 6.50 56 6.88 1.40 0.37 0.44 20 14 Crows Nest (18-10.65S;177-32.60E) 32.9 29.4 6.81 53.25 5.45 1.22 0.36 0.43 16 15 Korotogo Bridge (18-10.67S;177-32.59E) 27.3 28.9 3.47 45.9 5.68 5.82 0.47 1.14 11 16 Matai Kadavu Beach (18-10.77S;177-
31.05E) 31.8 30.1 5.81 36.76 2.97 1.69 0.63 0.35 15
17 Between Malevu & Vatukarasa (18-
11.17S;177-33.57E) 30.6 29.9 6.30 46.69 3.53 1.31 0.24 0.35 15
18 Warwick Hotel (18-13.69S;177-44.37E) 31.7 30.8 6.43 51.05 4.11 1.73 0.28 0.37 17
Mean 32.0 29.4 6.10 49.92 4.16 2.09 0.35 0.43 17 Mean Standard Error ±0.4 ±0.3 ±0.20 ±1.16 ±0.35 ±0.29 ±0.02 ±0.05 ±1 Standard Deviation 1.5 1.2 0.8 4.8 1.5 1.2 0.1 0.2 5
115
Table 13 which displayed a summary of Coral Coast ―baseline data‖ prior
to the monitoring period under this study (e.g. July 2005 to July 2006)
showed that salinity ranged from 27.2 ppt at Votua to 36.4 ppt for the
Fijian Resort shoreline with a mean of 33.8 ± 0.6 ppt. Water temperature
yielded a minimum of 21.2 ºC at Korotogo Bridge and a maximum of 29.6
ºC for Fijian Resort and Sovi Bay. Average temperature for the 17
monitored sites was 27.5 ± 0.5 ºC. Dissolved oxygen fluctuated between
2.29 mg/L for Korotogo Bridge and 8.3 mg/L for Tagaqe with an overall
mean of 6.94 ± 0.33 mg/L.
Faecal coliform levels showed a significant variance between sites. For
instance Fijian Resort yielded 1 count/100ml; Naviti Resort 31
count/100ml; Hideaway Resort 137 count/100ml; Votua Village 507
count/100ml; and Warwick Hotel 902 count/100ml. The average
coliform level for the Coral Coast sites was 169 ± 68 counts/100ml.
Nitrate (NO3-N) concentration was lowest at Matai Kadavu Beach with
0.10µmol/L and higher at Votua (3.34µmol/L) and Vatukarasa
(2.13µmol/L). Mean nitrate level was 1.11 ± 0.20 µmol/L. Site number 13
(i.e. Malevu Village) attained the least ammonia value of 1.39µmol/L
while Vatukarasa yielded the highest concentration of 9.36µmol/L
followed by Hideaway Resort at 7.93µmol/L. The average ammonia (NH3-
N) level for the Coral Coast prior to July 2005 was 5.24 ± 0.71 µmol/L.
116
The level of phosphate (PO4-P) was observed to be lowest at site number
7 (Sovi Bay) with 0.09µmol/L and varied between other sites, for
example: 0.28µmol/L for Matai Kadavu Beach; 0.59µmol/L for Fijian
Resort; 0.89µmol/L for Tabua Sands; 1.39µmol/L for Outrigger Resort
and 1.51µmol/L for Korotogo Bridge. Nevertheless the mean phosphate
concentration for the Coral Coast prior to July 2005 was 0.73 ± 0.10
µmol/L.
Total inorganic nitrogen to phosphorus ratio for the Coral Coast ranged
from 0.2 (site number 15) to a maximum of 13 at site number 8
(Hideaway Resort). However the mean nitrogen to phosphorus ratio for
the Coral Coast prior to July 2005 was 6 ± 1.
117
Table 13: Summarised “baseline” water quality data from the Coral Coast over a five year period prior to July 2005
(courtesy of Bale Tamata, Sarabjeet Singh, Luke Mosley & Bill Aalbersberg – IAS monitoring)
Site #
Place GPS location Sal (ppt)
Temp (ºC)
DO (mg/l)
F/Coliform (c/100ml)
NO3-N µmol/L
NH3-N µmol/L
PO4-P µmol/L
TIN:P ratio
1 Fijian Resort 18-08.62S; 177-25.76E 36.4 29.6 8.0 1 1.11 3.75 0.59 8 2 Outrigger 18-10.82S; 177-33.08E 34.5 27.9 6.54 45 1.78 2.37 1.39 3 3 Tubakula 18-10.86S; 177-33.46E 35.6 27.0 6.5 24 1.08 6.10 0.79 9 5 Votua Village 18-12.69S;177-42.89E 27.2 28.5 7.6 507 3.34 - 0.47 7 6 Tagaqe Village 18-11.91S; 177-39.75E 30.9 28.2 8.3 29 1.21 4.37 0.67 8 7 Sovi Bay 18-12.30S; 177-36.39E 31.7 29.6 7.54 7 0.35 - 0.09 4 8 Hideaway 18-11.92S; 177-39.32E 36.0 28.0 8.0 137 0.64 7.93 0.63 13 9 Naviti Resort 18-12.31S; 177-41.84E 33.6 28.8 7.4 31 0.63 3.98 0.71 7 10 Komave Village 18-13.38S; 177-45.71E 31.3 26.2 7.6 87 0.40 - 0.47 1 11 Tabua Sands 18-11.62S; 177-37.89E 35.8 28.6 7.9 4 1.08 6.71 0.89 9 12 Vatukarasa 18-10.85S; 177-36.22E 34.5 26.5 6.4 289 2.13 9.36 1.33 9 13 Malevu Village 18-10.85S; 177-33.62E 35.0 27.3 6.35 109 1.80 1.39 0.79 4 14 Crows Nest 18-10.65S; 177-32.60E 34.3 27.8 7.38 191 1.60 6.20 1.0 8 15 Korotogo Bridge 18-10.67S; 177-32.59E 33.9 21.2 2.29 - 0.25 - 1.51 0.2 16 Matai Kadavu
Beach 18-10.77S; 177-31.05E 34.0 28.7 5.98 - 0.10 - 0.28 0.4
17 Between Malevu & Vatukarasa
18-11.17S; 177-33.57E 34.1 27.4 7.35 - 0.71 - 0.20 3
18 Warwick Hotel 18-13.69S; 177-44.37E 36.1 26.9 6.8 902 0.65 5.51 0.59 10
Mean 33.8 27.5 6.94 169 1.11 5.24 0.73 6 Mean Standard Error ±0.6 ±0.5 ±0.33 ±68 ±0.20 ±0.71 ±0.10 ±1 Standard Deviation 2.4 1.9 1.4 253 0.8 2.4 0.4 4
118
6.5. Votua Creek water quality monitoring
Ongoing monitoring of water quality along the Coral Coast by the
Institute of Applied Science has shown that Votua Creek is one of the
relatively higher sources of pollution input into the coastal water due to
untreated wastewater and piggery farming. As a result there was need to
investigate the points of pollution from an upstream housing
development along the Votua Creek prior to an anticipated village
wetland system. Monitoring was undertaken from June to September
2006 (see Appendix D) and results are summarised in Table 14.
Data in Table 14 appear to show that faecal coliform levels generally
increased on a linear trend from the Votua water supply dam
downstream to the creek mouth. For instance Votua Dam had 140
counts/100ml; Upper Housing 165 counts/100ml; Lower Housing
yielded 535 counts/100ml; Votua Bridge 664 counts/100ml and Votua
Creek mouth 813 counts/100ml. In terms of drinking water supply, the
Votua Housing tap water had <1 coliform counts/100ml, 58
counts/100ml for Votua Village tap water and 33 counts/100ml for
Mike‘s Diver tap water. E/coli for creek water quality was lowest at the
Votua Dam site with 95 counts/100ml; followed by Upper Housing (135
counts/100ml); Lower Housing (296 counts/100ml) and the Creek
mouth (483 counts/100ml). The highest E/coli level was observed at the
Votua Bridge with 751 counts/100ml). Votua Housing tap water reached
119
<1 counts/100ml while the Village and Mike‘s Diver tap water yielded 48
and 21 counts/100ml respectively.
Conductivity also increased from the Votua Dam downstream to the
Creek mouth with a range between 0.11mS and 0.23mS. In terms of
water flow, it was generally ‗fast‘ at the Votua Dam and Upper Housing
sites; ‗medium‘ at Lower Housing; and ‗slow‘ at Votua Bridge and Creek
mouth. Votua Dam also obtained the lowest total suspended solids
reading with 8.3 mg/L followed by Upper Housing (9.7 mg/L), Lower
Housing (14.0 mg/L) and Creek Mouth (14.3 mg/L). Votua Bridge
displayed the highest suspended solids level of 15.3 mg/L. For biological
oxygen demand, the results showed no obvious variance between sites as
all monitored sites yielded <18 mg/L.
Ammonia (NH3-N) concentration for the Votua Creek ranged between
3.1µmol/L at the Dam site and 11.4µmol/L at the Bridge. Lower Housing
had the second highest ammonia level with 10.7µmol/L. Nitrate (NO3-N)
also showed a similar trend to ammonia at all the monitored sites along
the Votua Creek with a low of 1.96µmol/L at the dam to a high of
11.9µmol/L at the bridge. Nitrite (NO2-N) was least at the dam with
0.46µmol/L and increased downstream to a maximum of 1.12µmol/L at
the creek mouth. Total kjeldahl nitrogen (TKN) was least from the upper
creek dam site and relatively increased downstream to the creek mouth
120
ranging from 7.2 to 23.8 µmol/L. Total inorganic nitrogen (TIN) reached a
maximum at the bridge with 24.4µmol/L with the lowest concentration of
5.5µmol/L at the dam.
Phosphate (PO4-P) concentrations for the Votua Creek yielded
0.36µmol/L at the dam, exceeded upper housing with 0.39µmol/L,
0.78µmol/L for lower housing, Votua Bridge with 0.99µmol/L, and the
creek mouth with 1.36µmol/L. Total phosphorus also highlighted similar
trend as compared to phosphate with the dam site showing the lowest
value of 0.44µmol/L. Upper housing had 0.50µmol/L; lower housing
0.67µmol/L; 0.93µmol/L for Votua bridge whilst the creek mouth
attained 1.35µmol/L.
121
Table 14: A summary of Votua Creek mean water quality monitoring between June and September 2006 (n = 4)
Site FC
c/100ml
E.coli
c/100ml
Sal
ppt
Temp
ºC
Cond
mS
pH TSS
mg/l
Flow BOD
mg/l
NH3
µM
NO3
µM
NO2
µM
TIN
µM
TKN
µM
PO4
µM
TP
µM
Votua Dam
140
95
0.0
19.3
0.110
7.7
8.3
Fast
<18
3.1
1.96
0.46
5.5
7.2
0.36
0.44
Upper
Housing
165
135
0.0
18.1
0.111
7.4
9.7
Fast
<18
4.7
6.45
0.56
11.7
11.6
0.39
0.50
Lower
Housing
535
296
0.0
20.7
0.115
7.3
14.0
Med
<18
10.7
11.0
0.75
22.5
26.5
0.78
0.67
Votua
Bridge
664
751
0.2
20.9
0.124
7.3
15.3
Slow
<18
11.4
11.9
1.07
24.4
23.3
0.99
0.93
Votua
Creek
mouth
813
483
15.6
22.6
0.230
7.8
14.3
Slow
<18
10.3
10.7
1.12
22.1
23.8
1.36
1.35
Housing tap water
<1
<1
0.0
-
-
-
-
-
-
-
-
-
-
-
-
-
Village
tap water
58
48
0.0
-
-
-
-
-
-
-
-
-
-
-
-
-
Mike‘s Diver tap
water
33
21
0.0
-
-
-
-
-
-
-
-
-
-
-
-
-
122
6.6. Drum system experiment
6.6.1. Water quality monitoring program
Sampling and actual monitoring started in mid May until early June
2007 (Figure 39). For monitoring period 1 (i.e. large doses) which
comprised 3 doses per day; artificial greywater solutions were added
every 8am, 1pm and 5pm. After 3 days, samples were collected from
the 1pm dose. In regard to monitoring period 2 (e.g. moderate doses)
consisting of 6 doses per day; artificial greywater doses were added
every 8am, 10am, 12midday, 2pm, 4pm and 6pm. After 2 days,
samples were collected from the 12midday dose. Monitoring period 3
(small doses) included 12 doses per day, hence doses were added
every 8am, 8.40am, 9.20am, 10am, 10.40am, 11.20am, 12midday,
12.40pm, 1.20pm, 2pm, 2.40pm, and 3.20pm. After 1 day, samples
were collected for the 12midday dose. Results for the different dosage
regimes are summarised in Tables 15 to 17.
Figure 39: Sampling the ex-situ drum system experiment
123
Table 15: Summary of results for “Monitoring Period 1 – Large Doses”
Parameter Artificial
Greywater
Solution
High Loading
(HL)
Low Loading
(LL)
Temperature (ºC) 23.8 24.6 25.7 Salinity (ppt) 0.19 0.23 0.23
Conductivity (mS) 0.586 0.592 0.588
pH 6.3 7.6 8.2
Dissolved oxygen (mg/L) 1.96 2.33 2.31
Total Dissolved Solids (mg/L) 0.466 0.421 0.430
Faecal coliform (c/100ml) 1.2x106 1.7x105 2.2x105 E.coli (c/100ml) 5.0 x103 2.0x103 2.0x103
BOD (mg/L) 81 76 63
Total Suspended Solids (mg/L) 119 93 113
Total Kjeldahl Nitrogen (mg/L) 8.3 5.45 8.27
Phosphorus (mg/L) 9.2 8.33 7.61
During ―Monitoring Period 1 (i.e. large doses)‖ which is summarised in
Table 15; conductivity, dissolved oxygen, total dissolved solid and
E/coli showed very small variance between the High Loading and Low
Loading mesocosms. Conductivity was 0.592mS/cm for the high
Loading drum effluent and 0.588mS/cm for the low loaded system. A
dissolved oxygen level of 2.33 mg/L was obtained for high loading and
2.31 mg/L for low loading whilst E/coli levels remained constant at
2,000 counts/100ml for both mesocosms.
Faecal coliform concentration was relatively higher for the Low
Loading drum with 220,000 counts/100ml in comparison to 170,000
counts/100ml for the High Loading regime. Biological oxygen demand
was 76 mg/L for high loading and 63 mg/L for low loading. The low
loaded drum attained a higher total suspended solids value of 113
mg/L as opposed to 93 mg/L for the high loaded mesocosm. TKN
increased from 5.45 mg/L for high loading to 8.27 mg/L for the low
124
loading regime. Phosphorus decreased from 8.33 mg/L for high
loading to 7.6 mg/L for low loading.
Table 16: A summary of results for “Monitoring Period 2 – Moderate Doses”
Parameter Artificial Influent
Solution
High Loading
#1
High Loading
#2
High Loading
#3
Low Loading
#1
Low Loading
#2
Low Loading
#3
Temperature
(ºC)
26.4 25.6 25.6 25.7 26.3 26.3 25.7
Salinity (ppt)
0.23
0.27
0.28
0.28
0.28
0.33
0.30
Conductivity
(mS)
0.597
0.599
0.582
0.587
0.870
0.877
0.869
pH
5.7
7.8
7.8
7.9
8.1
8.3
8.3
DO (mg/L)
2.34
2.11
1.95
1.13
2.24
2.09
2.01
TDS (mg/L)
0.576
0.379
0.431
0.378
0.538
0.511
0.377
Faecal coliform
(c/100ml)
1.7x106
5.9x105
1.6x105
1.6x105
1.6x105
1.6x105
1.6x105
E.coli
(c/100ml)
2.0 x104
800
2.3x103
5.0x103
8.0x103
3.0x103
1.7x103
BOD (mg/L)
292
82
81
271
61
104
130
TSS (mg/L)
132
92
93
110
108
118
110
TKN (mg/L)
11.4
<2
<2
<2
9.94
6.09
<2
Phosphorus
(mg/L)
10.1
6.28
7.34
7.38
6.75
7.75
7.48
Note: #1 - an effluent sample collected immediately after dosing
#2 - a middle sample collected 30 minutes after the first sample
#3 - an end sample collected 60 minutes after the first sample
Table 16 depicting ―Monitoring Period 2 (e.g. moderate doses)‖ showed
lower conductivity levels of 0.582 to 0.599 mS/cm for the High
Loading drum effluent, while low loading ranged between 0.869 and
0.877 mS/cm. Salinity levels were slightly higher in the low loaded
drum than the high loaded system. The low loaded effluent was more
basic than the high loading drum as well. In regard to dissolved
oxygen, levels were higher for the low loaded drum ranging from 2.01
125
mg/L to 2.24 mg/L where as the high loading drum attained lower
values within 1.13 to 2.11 mg/L. Similarly, total suspended solids
(TSS) and total kjeldahl nitrogen (TKN) showed comparatively higher
concentrations for the low loaded drum as opposed to the high loading
mesocosm. Other parameters such as faecal coliform, E/coli, and
biological oxygen demand (BOD) displayed no obvious variance
between the two mesocosms.
For monitoring period 2, three periodic samples were collected from
each drum experiment in 30 minute intervals. Sample #1 was
referenced as the immediate effluent sample after dosing; sample #2
was a middle sample collected 30 minutes after the first sample; and
sample #3 was an end sample collected 60 minutes after the first
sample. From Table 16 it can be seen that the pH levels slightly rose
for both mesocosms from sample #1 to sample #3. On the contrary,
dissolved oxygen levels fell in the order from samples #1 to #3. For
E/coli there was an increase from sample #1 to sample #3 in the High
Loading drum whilst for the Low Loading drum there was reduction.
BOD, TSS and Phosphorus showed increases in the order from
samples #1 – 3 for both mesocosms. TKN was stable for the High
Loading experiment and decreased for Low Loading from sample #1 to
sample #3.
126
Table 17: Mean results for “Monitoring Period 3 – Small Doses”
Parameter Artificial
Influent
Solution
High
Loading
#1
High
Loading
#2
High
Loading
#3
Low
Loading
#1
Low
Loading
#2
Low
Loading
#3
Temp (ºC) 24.9 26.5 26.5 26.5 25.3 25.3 25.4
Sal (ppt)
0.14
0.29
0.28
0.28
0.33
0.30
0.29
Cond (mS)
0.557
0.696
0.581
0.583
0.885
0.624
0.606
pH
5.3
7.7
7.8
7.8
8.25
8.26
8.26
D/Oxygen
(mg/L)
2.81
1.65
1.35
1.21
2.23
2.11
1.36
TDS (mg/L)
0.618
0.451
0.378
0.379
0.573
0.405
0.393
Faecal
coliform
(c/100ml)
1.7x106
9.0x105
3.0x105
1.7x105
3.0x105
1.3x105
2.2x105
E.coli (c/100ml)
4.0 x104
6.0x103
4.0x103
4.0x103
2.6x104
8.0x103
1.4x104
BOD
(mg/L)
273
240
172
135
60
98
108
TSS (mg/L)
36
32
22
20
29
11
26
TKN (mg/L)
15.9
10.9
4.34
7.28
14.0
10.1
8.96
Phosphorus
(mg/L)
9.13
7.05
7.54
8.07
4.05
5.98
6.22
Note: #1 - an effluent sample collected immediately after dosing #2 - a middle sample collected 30 minutes after the first sample
#3 - an end sample collected 60 minutes after the first sample
Table 17 represents ―Monitoring Period 3 (i.e. small doses)‖. Effluent
from the low loading drum appears to be more basic than the effluent
from the high loading system. Dissolved oxygen was also higher for
the low loaded effluent than high loading. Faecal coliform, phosphorus
and biological oxygen demand concentrations for the high loading
design exceeded that of the low loading drum. However, E/coli and
total kjeldahl nitrogen levels were higher for the low loading drum
than the high loading regime.
127
In relation to the variance between 30 minute interval samples;
salinity, dissolved oxygen, conductivity, total dissolved solids, total
kjeldahl nitrogen and faecal coliform showed decreases from sample
#1 to sample #2 for both drums. For the high loaded drum obvious
reduction from sample #1 to #3 were observed for biological oxygen
demand and total suspended solids. Phosphorus and biological
oxygen demand were observed to rise in the low loading effluent from
sample #1 to sample #3.
6.6.2. Other monitoring
A gauge of the degree of soil-interface clogging was obtained by
comparing the hydrograph (i.e. effluent flow rate versus time) for the
two mesocosms following application of a single dose of 85 litres/dose
for the High Loading mesocosm and 42.5 litres/dose for the Low
Loading regime, both when first set-up (clean system) and at the end
of the experiment. The findings are displayed in Figure 40.
Figure 42: A graph showing the effluent flow rate vs
time for the two mesocosms
73
38
118
66
0
20
40
60
80
100
120
140
HL - 85 litres/dose LL - 42.5 litres/dose
Mesocosm regime
Tim
e (
min
ute
s)
Clean system
Clogged system
Figure 40 highlights that for a standard high loading dosage of 85
litres of artificial greywater solution, it took 73 minutes to elute when
the experiment was first set up (clean system) and 118 minutes to
Figure 40: A graph showing the effluent flow rate vs time for the two mesocosms
128
elute when the same dose was added at the end of the experiment.
The difference was 45 minutes. For a standard low loading dosage of
42.5 litres of artificial greywater solution, it took 38 minutes to elute
for the clean system and lasted 66 minutes for the clogged system at
the end of the experiment. The difference was 28 minutes.
In regards to a visual observation of the condition of the different
layers within the two mesocosms, it was found that signs of clogging
and organic matter build up were relatively higher within the ‗High
Loading‘ drum as compared to the ‗Low Loading‘ drum. Organic
matter build up was obvious on the coconut husk and shell layers and
decomposition was greatest at the deeper layers comprising the coral
rubble, sandy soil, and gravel producing an anoxic environment and
emitting septic-like foul smell. A display of some degree of clogging on
the coconut husk layer within the High Loading drum is highlighted in
Figure 41.
Figure 41: Some degree of clogging on the coconut husk layer within the High
Loading mesocosm
129
Chapter 7 Discussion
7.1. Tagaqe Village constructed wetland
Monitoring data for the Tagaqe Village constructed wetland which are
summarised in Table 10 showed that there was no significant
difference in temperature and salinity measurements for both the
untreated influent and treated effluent. For example the influent and
effluent temperatures were 28.6 ºC and 27.2 ºC respectively. Salinity
remained constant at zero ppt for both the untreated and treated
samples. These results are consistent with the warmer climate in Fiji
and the amount of daily household freshwater that enters the Tagaqe
wetland as blackwater and greywater. The constant and low salinity
levels are also related to the wetland pH levels of 6.80 (i.e. influent)
and 6.85 (effluent) which are quite close to a neutral pH value of 7.
Despite both wetland influent and effluent being slightly acidic, the
levels fall within ‗natural waters‘ characteristics of a pH range between
6.5 and 8.5. An acidic pH of 6.5 or lower may enhance odours in
anaerobic conditions (Mukhtar et al., 2004). According to Wood et al.,
(1999) temperature affects both the physical and biological activities
in wetland systems. Similarly, crucial nitrogen processes or nitrogen
fluxes depend on water chemistry and other wetland conditions, such
as climate, vegetation, water depth and water flow (Johnston, 1991;
Tanner, 2001a; Bastviken, 2006).
130
Dissolved oxygen concentrations showed slight improvement at the
outlet sampling point with 0.82 mg/L in comparison to 0.16 mg/L for
the untreated wetland influent. Dissolved oxygen is the oxygen freely
available in water which is vital to fish and other aquatic life and for
the prevention of odours. Therefore visual observation of wastewater
samples from the wetland (see Figure 40), which characterised treated
effluent as being clearer with an earthy odour rather than the clogged
and foul odour for untreated influent, can be partially attributed to
the observed improvement in dissolved oxygen. Research undertaken
on oxygen fluxes and ammonia removal efficiencies has found that
ammonia removal efficiency in constructed wetlands is often limited
by the amount of oxygen available in the system (Mei-Yin et al., 2001).
Electrical conductivity was observed to decrease significantly from the
untreated influent to treated effluent with a mean efficiency of around
83.3 percent. The influent yielded a mean of 0.12 mS/cm while 0.02
mS/cm was attained for the treated effluent. The reduction in
conductivity may be attributed to the decline in ions present (Tanner,
2004: personal communication).
Results indicated that the percentage removal or wetland treatment
efficiency for faecal coliform was 99.3 and E/coli 98.9 percent. The
treated wetland effluent contained a mean E/coli count of 15,000 per
100ml and faecal coliform count of 24,000 per 100ml. Both levels lie
within the range of discharge consent requirements issued for
131
constructed wetlands in New Zealand by Regional Councils, which
perceived the effectiveness of constructed wetlands by a discharge
faecal coliform range of 14 – 80,000 counts per 100ml (Tanner and
Sukias, 2002). High bacterial removal from the Tagaqe wetland
exceeding 98 percent for both faecal coliform and E/coli signify that
constructed wetlands, particularly in a warm tropical region is an
effective and affordable wastewater treatment alternative for coastal
villages and tourist hotels in terms of bacterial reduction and
disinfection.
In addition total suspended solids (TSS) were removed at 96.5 percent
from an average 886.6 mg/L in the inflow sample to 30.6 mg/L in the
treated effluent. Similarly, biological oxygen demand (BOD) decreased
from 324.8 (i.e. influent) to 17.3 mg/L for the treated effluent
correlating to 94.7 percent removal. The observed high BOD removal
efficiency for the Tagaqe wetland exceeds initial kinetic modelling
predictions of around 70 percent (Tanner, 2004: personal
communication). BOD is a measure of the amount of oxygen
consumed in the biological processes that break down organic matter
in water. Higher BOD implies a greater degree of pollution. The
significant rate of TSS and BOD removal is also higher than results
obtained by Tanner (2001b) for planted subsurface flow treatment
wetland systems which showed enhanced nitrogen and initial
phosphorus removal, but only small improvements in disinfection,
BOD and suspended solids removal. The trend may be attributed to
132
either wetland under-loading from the Tagaqe Chief‘s house or the
variance in climatic conditions between Fiji and New Zealand.
Other studies in France (Boutin et al., 1997), Switzerland (Schonborn
et al., 1997) and Thailand (Puetpaiboon and Yirong, 2004) also
reported higher removal efficiencies for TSS and BOD. For example,
investigations into reed bed filters for sewage treatment from small
communities in France yielded 92.5 percent for BOD and 94.5 percent
for TSS (Boutin et al., 1997) whilst constructed wetland research in
Switzerland obtained 95.8 percent for BOD removal (Schonborn et al.,
1997). A laboratory scale constructed wetland experiment in Thailand
achieved BOD and TSS removal efficiencies of 85 and 95 percent,
respectively (Puetpaiboon and Yirong, 2004). The performance of a
South Finger reedbed constructed wetland in the United Kingdom
indicated good treatment levels, with suspended solids reduction
around 80 percent and BOD generally above 60 percent (Worrall et al.,
1997). Another study on horizontal subsurface flow systems in the
German speaking countries (Geller, 1997) obtained very high
elimination rates far better than 90 percent for BOD. Tanner (1996)
also cited higher mean removals of 76-88 percent of suspended solids
and 77-91 percent of biological oxygen demand.
Nitrogen and Phosphorus fractions from the Tagaqe wetland seemed
to vary in their removal. Ammonia was removed at 82.6 percent while
nitrite attained 50 percent. A lower ammonia removal efficiency of 60
percent was found by Puetpaiboon and Yirong (2004) while Schonborn
133
et al. (1997) observed a higher ammonia reduction of 93.0 percent.
According to Keeney (1973), ammonia (NH3-N) formation occurs
through the mineralisation of organic matter under either anaerobic
or aerobic conditions. When ammonia combines with water within a
wetland it results in the formation of the ammonium ion (NH4+) which
can be absorbed by the plants and algae and converted back into
organic matter, or the ammonium ion can be immobilised onto
negatively charged soil particles (Mitsch and Gosselink, 1986).
Ammonium is eliminated when it is transformed to nitrate by the
bacterial process ‗nitrification‘ (Patrick and Reddy, 1976).
Nitrate did not show a decrease but instead increased from
1.04µmol/L for untreated influent to 6.08µmol/L for the treated
effluent. This equates to a rise of 484.6 percent. Despite other
research (US EPA, 1988; Drizo et al., 1997) highlighting reductions in
nitrate (NO3-N) within constructed wetland systems, Schonborn et al.,
(1997) reported similar NO3-N increases of 323.0 percent for a small
constructed wetland system in Switzerland which receives greywater
and liquid human waste. Nitrate is formed within a wetland system
when ammonium is transformed by the bacterial process called
‗nitrification‘. The process of nitrification (i) oxidises ammonium (from
the water column) to nitrite (NO2--N), and then (ii) nitrite is oxidised to
nitrate (NO3--N) (Keeney, 1973). Nitrate or nitrite is finally reduced to
gaseous end products, nitrous gas and dinitrogen gas, through the
bacterial process ‗denitrification‘ (Bastviken, 2006). Denitrification is
134
considered to be the predominant microbial process that modifies the
chemical composition of nitrogen in a wetland system and the major
process whereby elemental nitrogen is returned to the atmosphere
(Patrick and Reddy, 1976; Richardson et al., 1978; Johnston, 1991;
Vymazal, 2001; Trepel and Palmeri, 2002).
The observed concentration of nitrate from the Tagaqe wetland may
reflect the relative speed of transformation processes that occur within
the wetland, and is only marginally related to the initial nitrate
concentration. For instance, there was overall decline in the Total
Inorganic Nitrogen (TIN) concentration implying that the treated
effluent is mainly ammonia and unmineralised organic nitrogen. In
addition, the higher mean nitrate concentration at the wetland treated
sampling outlet may signify a delay in the ‗denitrification‘ process in
which nitrate is reduced to gaseous end products, N2O and N2, that
re-enter the atmosphere. Bandurski (1965) stated that ‗denitrification‘
occurs intensely in anaerobic environments. Results for the Tagaqe
wetland indicate that the treated effluent is less anaerobic as opposed
to the untreated influent. Hence, it may be reasonable that the
process of denitrification is hampered by slight aerobic conditions for
the wetland treated effluent.
Nevertheless there was enhanced total inorganic nitrogen (82.5
percent) and total kjeldahl nitrogen (75.5 percent) removal from the
wetland. Orthophosphate was eliminated by 75.5 percent and total
135
phosphorus 69.1 percent. These trends were compared to Schonborn
et al. (1997) who reported higher total nitrogen and total phosphorus
eliminations of 80 and 90.6 percent, respectively. Research in France
achieved total kjeldahl nitrogen removal of 76 percent whilst total
phosphorus removal was lower at 40 percent. Phosphate removal was
only 28 percent (Boutin et al., 1997). Another constructed wetland
case study in the United Kingdom attained an extremely high
phosphorus removal of 98-100 percent (Drizo et al., 1997). Tanner
(1996) also cited higher total phosphorus removal of 79-93 percent
while total nitrogen removed ranged from 65 to 92 percent.
Investigations into the effect of loading rate and planting on treatment
of dairy farm wastewaters in constructed wetlands reported 48 to 75
percent reduction of total nitrogen and 37 to 74 percent of total
phosphorus removal, in planted wetlands (Tanner et al., 1995).
Phosphate removal efficiency of the Tagaqe wetland decreased after a
period of time due to excess sediment adsorption and plant maturity.
This was initially predicted (Tanner, 2001; Headley, 2007: pers. com.).
Despite the high removal efficiencies within the constructed wetland,
there was no indication of the actual loading rate or residence time,
which can be attributed to lack of funding resulting in that limitation
of the experimental design.
7.2. Crusoe’s Resort wastewater treatment plant
Results for the Crusoe‘s Resort wastewater treatment plant (see Table
11) indicated that there was no variance in temperature and salinity
136
measurements for both the untreated influent and treated effluent.
Temperature measurements were 30.3 ºC for the influent and 30.7 ºC
for the effluent. As expected salinity levels remained at 0 ppt for
sampling points. Besides, pH levels of 6.76 for the untreated influent
and 6.90 for the effluent imply that the wastewater comprised ‗natural
water‘ characteristics of a pH range between 6.5 and 8.5 (Mukhtar et
al., 2004). Dissolved oxygen was significantly improved for the system
from 0.17 mg/L for the influent to 4.20 mg/L for the effluent. This
means that oxygen freely available in the effluent may be comparable
to sites that experience a high degree of pollution but can still support
other aquatic life (Clark, 2002). Conductivity also decreased
considerably from 0.74 mS/cm from the untreated influent to 0.02
mS/cm for treated effluent with a mean efficiency of 97.3 percent. The
reduction in conductivity is likely the effects of a decline in ions
present. This is reasonable with the predicted peak flow of 7,000
litres/day due to excess water usage in tourist hotels (Tanner, 2004:
personal communication; Innoflow, 2005).
In terms of bacteria elimination, the Crusoe‘s system recorded 93.5
percent for faecal coliform and 94.7 percent for E/coli. The coliform
values found in this study are similar to those found by Louden et al.,
(1985) for a recirculating filter system in Michigan (99.6 percent), and
to other studies in Maryland (99.5 percent; Piluk and Peters, 1994),
Oregon (96.7 percent; Ronayne et al., 1982), Quebec (97.3 percent;
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Roy and Dube, 1994), and Wisconsin (98 percent; Ayres Associates,
1998).
Total suspended solids and biological oxygen demand removals were
observed to be lower than the performance of the constructed wetland
at Tagaqe. Total suspended solids (TSS) yielded a removal of 63.6
percent while biological oxygen demand (BOD) was 71.7 percent. TSS
was eliminated from 35.2 mg/L for the influent to 12.8 mg/L for the
effluent, which satisfies the Innoflow expected treatment level of less
than 15 mg/L. BOD fell from 63.6 mg/L at the inflow to 18.0 mg/L at
the effluent, which is slightly above the predicted performance of less
than 15 mg/L (Innoflow, 2005).
TSS and BOD removal efficiencies attained for the Crusoe‘s treatment
plant seemed to be lower than other related studies. O‘Reilly et al.,
(2008) found TSS removal of up to 90.7 percent with a BOD reduction
range of 77.3-92.9 percent. Research in Michigan achieved 96 percent
removal for BOD and 85.7 percent for TSS (Louden et al., 1985); in
Maryland researchers reported removal efficiencies of 97.9 percent for
BOD and 89.3 percent for TSS (Piluk and Peters, 1994); whilst in
Oregon BOD was observed to be removed at 98.6 percent and TSS at
97.3 percent (Ronayne et al., 1982). Similar work undertaken in
Quebec by Roy and Dube (1994) cited 94.1 percent for BOD removal
and 96.1 percent for TSS; in Wisconsin (Ayres Associates, 1998) BOD
was eliminated at 98.3 percent and 98.4 percent for TSS; where as
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Owen & Bob (1991) obtained average BOD and TSS removal
efficiencies of 90.0 and 83.3 percent, respectively.
Moreover ammonia (NH3-N) was removed at 72.7 percent; nitrate
(NO3-N) 68.5 percent; and nitrite (NO2-N) 50.9 percent. Dissolved
inorganic nitrogen removal efficiency was 72.7 percent and total
kjeldahl nitrogen 50.1 percent. In addition, phosphate (PO4-P) removal
reached 70.7 percent while total phosphorus yielded 60.5 percent. The
nitrogen to phosphorus (N: P) ratio was 12 for untreated influent and
8 for treated effluent. The Crusoe‘s Resort system displayed direct
nitrate reduction of 68.5 percent as opposed to the Tagaqe
constructed wetland, which may imply that this treatment system is
more effective in achieving enhanced ‗nitrification‘ and ‗denitrification‘
processes. Mean N: P ratio for this system was 12 for the influent and
8 for the effluent which compares to a mean of 9 for the Tagaqe
wetland influent and 12 for the wetland effluent. This ratio gives an
indication as to whether a water or wastewater sample is enriched
with either N (ratio>20) or P (ratio<10) relative to unpolluted levels
(Mosley and Aalbersberg, 2003). There was considerable variability in
this ratio between the Tagaqe wetland and the Crusoe‘s Resort system
between the influent and effluent but it appears that the wastewater
within the Crusoe‘s system is more enriched with nitrogen at the
influent and phosphorus at the effluent.
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Generally nitrogen and phosphorus removal efficiency at the Crusoe‘s
Resort system were less than 73 percent and did not meet the initial
predicted nutrient removal of more than 90 percent. However,
dissolved inorganic nitrogen removal of around 72.7 percent is
relatively similar to the expected system transformation of 75 percent
of nitrogen fractions into nitrogen gaseous end products, and only 15
percent would be disposed into flower gardens and eventually seeping
into the coral reef areas (Shortt, April 2005: personal communication).
Nutrient removal efficiency from similar systems elsewhere also
showed significant variability. For example, a study in Ireland for a
related system indicated total inorganic nitrogen (TIN) removal of 40-
81.6 percent while ammonia achieved 89.8 percent (O‘Reilly et al.,
2008). Research in Michigan cited total kjeldahl nitrogen (TKN)
elimination of 95.8 percent and TIN reduction of 52.7 percent (Louden
et al., 1985). A TKN removal efficiency of 98.1 percent and 45.2
percent of TIN were observed by Ronayne et al., (1982). Roy and Dube
(1994) found 79.1 percent for TKN and 46.7 percent for TIN while a
study in Wisconsin by the Ayres Associates (1998) reported 95.5
percent removal for TKN and 75.7 percent for TIN. Other studies
showed >95 percent of TKN removal (Owen and Bob, 1991) and 64.9
percent of TIN removal (Piluk and Peters, 1994).
In comparing the Tagaqe wetland with the Crusoe‘s system, it was
noted in the results obtained that the influent to the Crusoe device
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was much cleaner than the influent to Tagaqe. Therefore, it was more
difficult for the Crusoe system to achieve a high removal efficiency.
The cleaner influent at Crusoe is probably a reflection of the more
complex and larger primary treatment via at least four separate septic
tanks before being passed through the system.
7.3. Coral Coast water quality monitoring
Water quality data from the Coral Coast sites monitored in this study
(i.e. Table 12) showed that short term temporal differences were the
major source of variation for salinity, temperature, dissolved oxygen
and conductivity for the 17 sites monitored. For instance, mean
dissolved oxygen and temperature were greater over sampling times
consistent with increases in solar irradiation, and surface water
mixing due to strengthening wind conditions (e.g. sea breezes) whilst
salinity and conductivity were influenced by freshwater or seawater
flushing and dilution effects (Clark, 2002; Castro and Huber, 2003).
Generally, salinity varied between 27.3 ppt (parts per thousand) at
Korotogo Bridge and 34.1 ppt at Malevu Village with an average of
32.0 ± 0.4 ppt. The low salinity measurement at Korotogo Bridge is
probably due to freshwater input from a creek running under the
bridge. The addition of masses of freshwater from rivers and creeks
along the Coral Coast may have some influence on this result. Water
temperature ranged from 27.2 ºC for Tabua Sands and Komave to
32.1 ºC for Naviti Resort shoreline with a mean of 29.4 ± 0.3 ºC. The
mean temperature is within the range of normal ocean temperature,
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which is from 20 to 30 ºC (ANZECC, 2000). Overall, there was no
significant difference in salinity and temperature ranges for the Coral
Coast sites in this study as compared to baseline monitoring data
obtained prior to July 2005 by the Institute of Applied Sciences (see
Table 13).
Dissovled oxygen (DO) levels were relatively moderate with a minimum
of 3.47 mg/L at Korotogo Bridge and a maximum of 7.13 mg/L at
Outrigger Resort shoreline. The mean DO level for the 17 sites was
6.10 ± 0.20 mg/L. Apparently the mean DO value attained in this
study is slightly lower than the level observed for a Coral Coast
baseline monitoring data which yielded an average DO of 6.94 ± 0.33
mg/L for the same sites. However, both values fall within the
recommended DO standard of >6 mg/L for nearshore waters to
support coral reefs and recreation in Australia and New Zealand
(ANZECC, 2000).
Despite this study not investigating the coliform pollution along the
Coral Coast, baseline monitoring data indicated that there was a
significant variance between sites. For instance Fijian Resort yielded 1
count/100ml; Naviti Resort 31 count/100ml; Hideaway Resort 137
count/100ml; Votua Village 507 count/100ml; and Warwick Hotel
902 count/100ml. The average coliform level for the Coral Coast sites
was 169 ± 68 counts/100ml (IAS Monitoring Data, 2005:
unpublished). Considering the average coliform level of 169
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counts/100ml coupled with the lower end of the standard error of ±
68 counts/100ml, it can be assumed that faecal coliform levels at the
Coral Coast is within the recommended standard for nearshore waters
to support coral reefs and recreation in Australia and New Zealand,
which is <150 counts/100ml (ANZECC, 2000).
In reference to nutrients, Komave and Tagaqe shorelines yielded the
lowest nitrate (NO3-N) values of 2.33µmol/L and 2.58µmol/L,
respectively. The highest nitrate concentrations were observed for
Malevu, Outrigger Resort and Tubakula Resort shorelines with levels
of 6.88µmol/L; 6.68µmol/L; and 6.18µmol/L respectively. The average
nitrate level for the 17 monitored sites was 4.16 ± 0.35µmol/L. Site
number 17 (e.g. between Malevu & Vatukarasa) attained the lowest
nitrite (NO2-N) concentration of 0.24µmol/L and the highest was
observed at site number 16 (Matai Kadavu Beach) with 0.63µmol/L.
The average nitrite value for the Coral Coast was 0.35 ± 0.02 µmol/L.
Baseline monitoring data (IAS Monitoring Data, 2005: unpublished)
obtained prior to the commencement of this study in July 2005
showed nitrate (NO3-N) concentration was lowest at Matai Kadavu
Beach with 0.10µmol/L and higher at Votua (3.34µmol/L) and
Vatukarasa (2.13µmol/L). Mean nitrate level was 1.11 ± 0.20 µmol/L.
A comparative analysis appears to show an increase in mean nitrate
concentration on the Coral Coast from 1.11 ± 0.20µmol/L to 4.16 ±
0.35µmol/L between 2005 and 2006. Nitrite concentration is often
143
regarded as insignificant to nitrate and ammonia fractions but nitrite
levels for unpolluted waters often varied from 0 – 0.22 mol/L (Wetzel,
1975). Another study on an unpolluted lagoon yielded nitrite
concentrations ranging from 0.05 – 0.24 mol/L (Yamamuro et al.,
1991). Therefore the mean nitrite level found in this study (0.35 ±
0.02µmol/L) is above results obtained for unpolluted waters, but is
still below a mean nitrate value of 0.59µmol/L reported for the Port of
Suva (Tamata et al., 1992).
The mean nitrate level in this study (4.16 ± 0.35µmol/L) exceeds
recommended nitrate and nitrite standard range for nearshore waters
to support coral reefs and recreation in Australia and New Zealand,
which is between 0.14-0.57 µmol/L (ANZECC, 2000); as well as
critical nitrogen levels considered healthy for coral reefs without being
overgrown by algae which is 1.0 mol/L of nitrogen (N) as nitrate,
ammonia or nitrite (Bell et al., 1987; Bell, 1992; Goreau and Thacker,
1994). However higher nitrate values up to 4.8µmol/L (Naidu et al.,
1991); 7.01mol/L (Mosley and Aalbersberg, 2003); 27.05µmol/L
(Singh and Mosley, 2004: unpublished); 98.57µmol/L (Tamata et al.,
1992); and 357.1 µmol/L (Naidu and Morrison, 1988) were also
reported for other studies on the Coral Coast, Laucala Bay, Suva
Harbour and elsewhere in Fiji. Research on the highly populated
nearshore waters of Suva Harbour and Laucala Bay by Naidu &
Morrison (1988) and Naidu et al., (1991) yielded a mean nitrate level of
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17.29mol/L. Similar study by Tamata et al., (1992) found a mean
nitrate value of 5.47mol/L. An unpublished assessment of nutrient
status in Laucala Bay from 2003 – 2004 found an average nitrate level
of 1.77 µmol/L (Singh and Mosley, unpublished). Another
unpublished report of nutrient pollution in the Laucala Bay and Suva
Harbour in 2004 yielded a mean nitrate concentration of 3.68µmol/L
(Taloiburi, unpublished).
Ammonia (NH3-N) results in this study varied between 0.98µmol/L at
Tabua Sands and 5.82µmol/L at Korotogo Bridge with a mean of 2.09
± 0.29 µmol/L. On the other hand, baseline monitoring data prior to
2005 showed that site number 13 (i.e. Malevu Village) attained the
least ammonia value of 1.39µmol/L while Vatukarasa yielded the
highest concentration of 9.36µmol/L followed by Hideaway Resort at
7.93µmol/L. The average ammonia (NH3-N) level for the Coral Coast
prior to 2005 was 5.24 ± 0.71 µmol/L. According to Tanner and Gold
(2004) coastal sites that attain higher nutrient levels are likely to be
the effects of piggery and/or improper treated wastewater. Comparing
this study with the baseline data, it is obvious that there was a
general decrease in the mean ammonia level from 5.24 ± 0.71 µmol/L
before 2005 to 2.09 ± 0.29 µmol/L by 2006. For the Astrolabe lagoon
where there is insignificant pollution, levels of ammonia obtained were
in the range 0.05 – 0.20 mol/L (Yamamuro et al., 1991). Ammonia is
145
a better indicator for sewage pollution and anaerobic conditions
compared to nitrate and nitrite (Hawker and Connell, 1992).
Phosphate (PO4-P) concentration was least at Hideaway Resort
(0.28µmol/L) and Fijian Resort (0.31µmol/L) shorelines whilst Votua,
Vatukarasa and Korotogo Bridge yielded the highest phosphate levels
of 0.51; 0.58; and 1.14 µmol/L, respectively. A mean PO4-P of 0.43 ±
0.05 µmol/L was obtained for the Coral Coast water quality
monitoring. For the baseline monitoring data prior to 2005 (IAS
Monitoring Data, unpublished), phosphate was observed to be lowest
at site number 7 (Sovi Bay) with 0.09µmol/L and varied between other
sites. For example: 0.28µmol/L for Matai Kadavu Beach; 0.59µmol/L
for Fijian Resort; 0.89µmol/L for Tabua Sands; 1.39µmol/L for
Outrigger Resort and 1.51µmol/L for Korotogo Bridge. Mean
phosphate concentration for the Coral Coast prior to 2005 was 0.73 ±
0.10 µmol/L. Comparatively the average phosphate level for the Coral
Coast declined from 0.73 ± 0.10 µmol/L before 2005 to 0.43 ± 0.05
µmol/L by 2006. However both mean values exceed the critical
phosphorus level considered healthy for coral reefs without being
overgrown by algae, which is 0.1mol/L of phosphorus as
orthophosphate and organophosphate (Bell et al., 1987; Bell, 1992;
Goreau and Thacker, 1994). Similarly the mean phosphate
concentration in this study is higher than the recommended level
0.16mol/L for nearshore waters in Australia and New Zealand
(ANZECC, 2000).
146
Other studies (Crossland and Barnes, 1983) observed phosphate
levels suitable for normal coral growth to be within the range of 0.11 -
0.32 mol/L whilst phosphate levels as high as 0.74mol/L had been
reported from studies of Australian fringing reefs (Blake and Johnson,
1988). In addition, phosphate values obtained for the Astrolabe lagoon
seagrass bed was lower at 0.08 – 0.15 mol/L (Yamamuro et al., 1991)
with an average phosphate concentration of 0.07mol/L (Morrison et
al., 1992).
Research on moderate polluted coastal waters along the Coral Coast
in Fiji (Mosley and Aalbersberg, 2003) found phosphate levels that
exceeded thresholds considered harmful to coral reef ecosystems.
Phosphate levels for seawater varied between 0.07 – 1.51 mol/L with
an average of 0.21 mol/L. For freshwater samples, phosphate
concentrations ranged from 0.50 – 3.40 mol/L with a mean of 1.30
mol/L. An assessment of nutrient status in Laucala Bay from 2003 –
2004 found phosphate levels ranging between 0.46 – 11.01 µmol/L
with a mean of 0.95 µmol/L (Singh and Mosley, unpublished). Another
unpublished report of nutrient pollution in the Laucala Bay and Suva
Harbour in 2004 yielded a mean phosphate level of 1.20µmol/L
(Taloiburi, unpublished).
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The nitrogen to phosphorus (N: P) ratio for this study ranged between
9 and 29 with an average of 17 ± 1. This is significantly higher than
the mean N: P ratio of 6 ± 1 observed for the baseline monitoring data
for similar sites along the Coral Coast prior to 2005 (IAS Monitoring
Data, unpublished). Water quality research by Morrison et al., (1992)
for other unpolluted sites in Fiji yielded a mean N: P ratio of 10 while
Mosley and Aalbersberg (2003) reported a mean N: P ratio of 8 for the
same sites monitored in this present study along the Coral Coast. In
open ocean seawater production is thought of as being influenced by
the mole ratio of concentrations of nitrogen to phosphorus in the
water. For average seawater the nitrogen (N) to phosphorus (P) mole
ratio is about 15 N: 1 P which reflects the ratio of their utilisation by
phytoplankton (Collier, 1970). The N: P ratio provides an indication as
to whether a water sample is enriched with either N (ratio>20) or P
(ratio<10) relative to unpolluted levels (Mosley and Aalbersberg, 2003).
There was considerable variability in the N: P ratio between different
sites in this study, but in general the seawater within the fringing reef
on the Coral Coast is similar to the N: P ratio (i.e. 15) in open ocean
seawater yet slightly enriched with nitrogen than phosphorus. Efforts
had been made to encourage hotels to switch to non Phosphorus
detergents in 2004 and this may be reflected in these results.
7.4. Votua Creek water quality monitoring
Votua Creek was monitored from June to September 2006 in order to
attain a clear understanding of the major sources of wastewater
pollution along the creek as part of preliminary investigations toward
148
a village wetland system initiative, coordinated by the Institute of
Applied Sciences in partnership with the National Institute of Water
and Atmospheric Research (NIWA) in New Zealand.
Results (Table 14) showed that there was 5-6 fold increase in faecal
coliform from the dam and above housing down to the creek mouth.
Faecal coliform levels generally increased on a linear trend from
140counts/100ml at the Votua dam to 813counts/100ml at the
downstream creek mouth. Similarly, E/coli was least at the Votua
dam with 95counts/100ml; followed by upper housing (135
counts/100ml); lower housing (296 counts/100ml); creek mouth (483
counts/100ml); and the Votua bridge with 751 counts/100ml. This
correlates to a 5-8 fold increase from the dam and downstream.
Unfortunately both the faecal coliform and E/coli levels observed at
the dam were very high for a human drinking water supply without
further treatment (Tanner, 2006: personal communication; WHO,
2004). The mean upper housing faecal coliform measurements of 165
counts/100ml and 135 counts/100ml of E.coli were within
recommended nearshore standards for bathing and recreation, which
is 150-200 counts/100ml (ANZECC, 2000). However other common
bathing sites at the lower housing vicinity and the bridge indicated
very unsafe coliform levels.
Presence of coliform in the water body can be an indicator of sewage
discharge into the water either of mammalian or avian origin. Many of
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the faecal coliform bacteria in human waste are harmless. However
there are disease organisms or pathogens than can cause harm to
human health. These include bacteria such as typhoid, or viruses
such as hepatitis B. Direct contact with these pathogens or pollution
of the water supply can result in infections. This poses a public health
risk (nausea, vomiting, diarrhoea, ear, and throat infections) to people
who use the waters for bathing (New Zealand Ministry of
Environment, 2003). Another likely effect of direct contact with
coliform polluted water is skin irritation and scratchiness (Clark,
2002). Therefore, it is important that levels of faecal coliform do not
exceed recreational exposure standards.
In terms of drinking water supply, less than 1 count/100ml of faecal
coliform and E/coli were observed for the Votua Housing tap water.
The Votua Village drinking tap water reached 58 counts/100ml of
faecal coliform and 48 counts/100ml of E.coli. A nearby diving
centre‘s (i.e. Mike‘s Divers) drinking tap water obtained 33 and 21
counts/100ml of faecal coliform and E.coli, respectively. According to
World Health Organisation drinking water quality standards of <1
count/100ml (WHO, 2004) only the Votua Housing tap water is safe
for usage, but the village and Mike‘s Divers tap water are unsafe as
drinking water without further treatment. This is not surprising as the
village and diving centre drinking water was channelled directly from
the dam without any disinfection. The housing drinking water is
sourced near the upper housing sampled site and pumped up into a
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chlorine treated storage tank on the ridge before distribution to the
households.
Conductivity increased from the dam to the creek mouth with a range
between 0.11mS and 0.23mS. In terms of mean stream flow it was
generally ‗fast‘ at the dam and upper housing; ‗medium‘ at lower
housing; and ‗slow‘ at the bridge and creek mouth. Total suspended
solids at the dam was 8.3 mg/L but elevated to the creek mouth with
14.3 mg/L. Votua Bridge displayed the highest suspended solids level
of 15.3 mg/L. The stream flow and relative suspended solids trend
may imply that contamination downstream was likely due to storm
flows from the upper stream pollution points including human and
animal (i.e. piggery) wastewater. Results for biological oxygen demand
(BOD) showed no variance between sites as all monitored locations
yielded <18 mg/L. This is likely due to the minimum detection level of
18 mg/L for the APHA5210B Method used by the Institute of Applied
Science laboratory to determine BOD (APHA, 2005). In addition most
of the sampling schedules on the Votua Creek were undertaken
during rising tides so dilution may have an influence on the creek
mouth data.
Ammonia concentration ranged from 3.1µmol/L at the dam to
11.4µmol/L at the Bridge. Nitrate also showed a similar trend to
ammonia at all the monitored sites along the Votua Creek with a low
of 1.96µmol/L at the dam to a high of 11.9µmol/L at the bridge.
151
Nitrite was least at the dam with 0.46µmol/L and increased
downstream to a maximum of 1.12µmol/L at the creek mouth. Total
kjeldahl nitrogen was least from the dam and relatively increased to
the creek mouth ranging from 7.2 to 23.8 µmol/L. Total inorganic
nitrogen reached a maximum at the bridge with 24.4µmol/L with the
lowest concentration of 5.5µmol/L at the dam. River monitoring
results obtained by Mosley and Aalbersberg (2003) along the Coral
Coast also showed higher nitrate values that ranged from 1.9 to 24.7
µmol/L. Generally the critical nitrogen concentration considered
healthy to be deposited into the nearshore coastal waters without
affecting coral reefs is 1.0 mol/L of nitrogen as nitrate, ammonia or
nitrite (Bell et al., 1987; Bell, 1992; Goreau and Thacker, 1994). The
nitrogen (e.g. ammonia, nitrate and nitrite) levels observed at the
bridge and creek mouth significantly exceeds recommended
standards.
Phosphate concentrations for the Votua Creek were 0.36µmol/L at the
dam, followed by upper housing with 0.39µmol/L, 0.78µmol/L for
lower housing, Votua bridge with 0.99µmol/L, and the creek mouth
with 1.36µmol/L. Total phosphorus also showed similar trend with
the dam showing the lowest value of 0.44µmol/L. Upper housing had
0.50µmol/L; lower housing 0.67µmol/L; 0.93µmol/L for Votua bridge
whilst the creek mouth attained 1.35µmol/L. Earlier research on
selected river water along the Coral Coast by Mosley and Aalbersberg
(2003) reported phosphate values ranging from 0.50 to 3.40 µmol/L.
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The critical phosphorus level considered healthy for coral reefs
without being overgrown by algae is 0.1mol/L of phosphorus as
orthophosphate and organophosphate (Bell et al., 1987; Bell, 1992;
Goreau and Thacker, 1994). Therefore, phosphate and phosphorus
levels observed at the creek mouth are higher than normal accepted
standards.
7.5. Drum system experiment
Greywater is the wastewater generated from showers, bathtubs, hand
basins, laundry, washing machines and kitchen sinks and
consequently contains a mixture of soaps, detergents, food particles,
fats, oils, soil, hair, and potentially some small amounts of faecal
matter and urine. Studies (Urban Water Research Association of
Australia, 1996; Jefferson et al., 1999 & 2001; Eriksson et al., 2002;
Brown and Palmer, 2002; Toifl et al., 2006) suggest that greywater has
a similar organic strength to domestic wastewater, but relatively low
suspended solids (i.e. greater proportions of the contaminants are
dissolved).
7.5.1. Monitoring period 1 (large doses)
During ―monitoring period 1‖ whereby the mesocosms were loaded
with three large doses per day between 8am to 5pm (i.e. working
hours), conductivity, dissolved oxygen, total dissolved solid and E/coli
showed very small variance between the High Loading and Low
Loading mesocosms. In the ex-situ experiment, the High Loading drum
represents ‗one‘ onsite drum system per household at Votua Village
while the Low Loading drum correlates to ‗two‘ onsite drum per
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household (Tanner and Headley, 2007: personal communication).
Therefore assuming that households dispose only three large periodic
doses per day (e.g. morning, midday, evening) into the onsite drum
greywater treatment systems, there would be less variance between
the one drum per household (i.e. High Loading drum) and the two
drum per household (i.e. Low Loading drum) in how they treat
conductivity, dissolved oxygen, total dissolved solid and E/coli.
The faecal coliform concentration was relatively higher for the Low
Loading drum with 220,000 counts/100ml in comparison to 170,000
counts/100ml for the High Loading regime. The same trend was
observed for total suspended solids where the low loaded drum
attained 113 mg/L as opposed to 93 mg/L for the high loaded
mesocosm. Similarly, total kjeldahl nitrogen yielded 5.45 mg/L for
high loading and 8.27 mg/L for the low loading regime. These results
may imply that for three large periodic doses per day (e.g. morning,
midday, evening) into the onsite drum greywater treatment systems at
Votua Village, the one drum per household would be more efficient
than two drum per household in terms of faecal coliform, total
suspended solids and total kjeldahl nitrogen treatment.
On the other hand, biological oxygen demand measured 76 mg/L for
high loading and 63 mg/L for low loading. Phosphorus was 8.33 mg/L
for high loading and 7.6 mg/L for low loading. This indicates that for
phosphorus and biological oxygen demand treatment, the two drum
154
per household would be more efficient than one drum per household,
assuming a discharge of three large periodic doses per day.
Nevertheless both ex-situ mesocosms (e.g. low loading and high
loading) showed some degree of removal efficiencies for monitored
parameters from the prepared artificial greywater solution that was
passed through the experiment, but removal efficiencies varied
between parameters and the loading regimes.
7.5.2. Monitoring period 2 (moderate doses)
―Monitoring period 2‖ was reflective of six moderate doses per day
between 8am and 6pm (i.e. period greywater discharges every two
hours). Results indicated lower conductivity levels of 0.582 to 0.599
mS for the High Loading drum effluent, while low loading ranged
between 0.869 and 0.877 mS/cm. Salinity levels were slightly higher
in the low loaded drum than the high loaded system. The low loaded
effluent recorded a higher pH than the high loading drum as well. In
regard to dissolved oxygen levels, the values were higher for the low
loaded drum ranging from 2.01 mg/L to 2.24 mg/L where as the high
loading drum attained lower values within 1.13 to 2.11 mg/L.
Assuming that household greywater discharges at Votua Village is
achieved by moderate flushes every two hours (e.g. 8am, 10am,
12midday, 2pm, 4pm and 6pm), then the two drum per household
(i.e. low loading drum) would be the better set up to improve salinity,
dissolved oxygen and pH as opposed to one drum per household (i.e.
high loading).
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Total suspended solids and total kjeldahl nitrogen showed
comparatively higher concentrations for the low loaded drum as
opposed to the high loading mesocosm. Other parameters such as
faecal coliform, E/coli, phosphorus and biological oxygen demand
displayed no obvious variance between the two mesocosms. This
signifies that for total suspended solids (TSS) and total kjeldahl
nitrogen (TKN) treatment, the one drum per household (i.e. high
loading) would be the ideal system choice. However, the choice of
either one drum per household (high loading) or two drum per
household (low loading drum) in Votua does not necessarily matter for
elimination of other crucial water quality parameters including faecal
coliform, E/coli, phosphorus and biological oxygen demand. Both ex-
situ drum mesocosms displayed variable removal efficiencies for
monitored parameters from the prepared artificial greywater solution
that was passed through the experiment.
Samples were collected immediately after dosing (#1), 30 minutes after
the first sample collection (#2), and 60 minutes after the first sample
collection (#3) (Tanner and Aalbersberg, 2007: personal
communication). In terms of variance between samples which were
collected 30 minutes apart from the ex-situ drum experiment; pH,
BOD, TSS and phosphorus were observed to increase for both
mesocosms at the end sample. This means that for both mesocosms,
removal efficiency was greatest at the immediate sample (#1) for BOD,
TSS and phosphorus, except pH. Dissolved oxygen removal efficiency
156
was also enhanced at the immediate sample and deteriorated at the
end sample for both mesocosms. Similarly, E/coli removal efficiency
for the high loading drum was higher in the immediate sample and
least in the end sample whilst for the low loading drum; elimination
was elevated at the end sample. TKN was stable for the high loading
experiment and decreased for low loading from sample #1 to sample
#3, which implies that TKN removal efficiency for the low loading
drum was highest at the end sample.
7.5.3. Monitoring period 3 (small doses)
―Monitoring period 3‖ comprised 12 small doses per day periodically
added every forty minutes (i.e. 8am, 8.40am, 9.20am, 10am, 10.40am,
11.20am, 12midday, 12.40pm, 1.20pm, 2pm, 2.40pm, and 3.20pm).
Results indicated that the pH from the low loading drum appears to
be higher than the effluent from the high loading system. Dissolved
oxygen was also higher for the low loaded effluent than high loading.
Faecal coliform, phosphorus and biological oxygen demand
concentrations for the high loading design exceeded that of the low
loading drum. However, E/coli and total kjeldahl nitrogen levels were
higher for the low loading drum than the high loading regime. These
findings imply that by assuming the greywater discharge from Votua
households to be emitted every forty minutes in an average of 12
small doses per day; then the two drum per household (i.e. low
loading drum) set up would be more efficient in improving dissolved
oxygen, pH, faecal coliform, phosphorus and biological oxygen
demand whilst the one drum per household (i.e. high loading drum)
157
set up would enhance total kjeldahl nitrogen and E.coli removal.
Nevertheless both systems would achieve general greywater removal
efficiencies, despite possible variances between parameters.
In relation to the variance between 30 minute interval samples; total
kjeldahl nitrogen and faecal coliform elimination efficiencies were
achieved at the middle sample (#2) for both mesocosms. Biological
oxygen demand and total suspended solids achieved maximum
treatment efficiency at the end sample (#3) for the high loaded drum.
In the low loading drum, phosphorus and biological oxygen demand
removal were enhanced at the immediate sample (#1).
7.5.4. Other monitoring
The most likely mode of long-term failure of this type of greywater
management system was clogging or blockage of the soil surface
where the partially treated greywater infiltrates into the natural soil. A
―bio-mat‖ of biofilm, slime and accumulated organic solids was
attributed to be the main cause of such clogging (Urban Water
Research Association of Australia, 1996; Toifl et al., 2006). A gauge of
the degree of soil-interface clogging was obtained by comparing the
hydrograph (i.e. effluent flow rate versus time) for the two mesocosms
following application of a single dose, both when first set-up (i.e. clean
system) and at the end of the experiment (Headley, 2007: personal
communication). Monitoring data showed that for a standard high
loading dosage volume of 85 litres of artificial greywater solution, it
took 73 minutes to elute in the clean system and 118 minutes for the
158
same volume at the end of the experiment. The difference was 45
minutes. For a standard low loading dosage volume of 42.5 litres of
artificial greywater solution, it took 38 minutes to elute for the clean
system and lasted 66 minutes for the clogged system at the end of the
experiment. The difference was 28 minutes. This supports initial
prediction which anticipated the time taken for a dose to drain
through the system to increase as the soil-interface becomes clogged
and the degree of clogging to be greatest in the highly loaded
mesocosm (Headley, 2007: personal communication).
159
Chapter 8 Conclusions
In conclusion it is undeniable that most tourist hotels and villages on
the Coral Coast of Fiji are situated along the coastline. As a result of
the extreme pressure of anthropogenic actions exacerbated by animal
wastes and natural phenomenon, the abundance of seaweed growth
indicating enriched nutrient levels in nearshore coral reef areas and
higher coliform levels signifying sewage presence were observed
recently along the Coral Coast. As a result, sanitation experts
recommended various wastewater treatment initiatives which resulted
in the implementation of a constructed wetland at Tagaqe Village and
a commercial wastewater treatment system at Crusoe‘s Resort beside
other initiatives that were not monitored under this study. Monitoring
was undertaken on the planted gravel bed constructed wetland at
Tagaqe Village, Crusoe‘s Resort wastewater treatment system, Coral
Coast nearshore sites, Votua Village Creek, and an ex-situ greywater
treatment drum experiment.
Results for the wetland showed removal efficiency range of 94.7-99.3
percent for faecal coliform, E/coli, total suspended solids (TSS) and
biological oxygen demand (BOD). Nitrogen elimination ranged between
50 percent for nitrite and 82.6 percent for ammonia. Total kjeldahl
nitrogen (TKN) achieved 75.5 percent. Total phosphorus and
phosphate attained 69.1 and 75.5 percent, respectively.
160
The wastewater treatment system at Crusoe‘s Resort indicated
removal range of 63.6-94.7 percent for faecal coliform, E/coli, TSS
and BOD. Nitrite yielded 50.9 percent; nitrate 68.5 percent; ammonia
72.7 percent; and TKN 50.1 percent. Total phosphorus reached 60.5
percent while phosphate obtained 70.7 percent.
For the Coral Coast nearshore water quality, results showed an
average salinity of 32.0 ± 0.4 ppt; temperature of 29.4 ± 0.3 ºC;
dissolved oxygen level of 6.10 ± 0.20 mg/L; and mean conductivity
level of 49.92 ± 1.16 mS/cm. In terms of nutrients, the average nitrate
concentration of nearshore waters along the Coral Coast was 4.16 ±
0.35µmol/L. Ammonia yielded a mean value of 2.09 ± 0.29µmol/L
while nitrite attained 0.35 ± 0.02µmol/L. The mean phosphate level
for the Coral Coast was 0.43 ± 0.05 µmol/L. The N: P ratio for the
Coral Coast waters in this present study was 17 ± 1.
Results for Votua Creek showed a 5-6 fold increase in faecal coliform
from the dam and above housing down to the creek mouth. Faecal
coliform levels generally increased on a linear trend from
140counts/100ml at the Votua dam to 813counts/100ml at the
downstream creek mouth. Similarly, E/coli was least at the Votua
dam with 95counts/100ml; followed by upper housing (135
counts/100ml); lower housing (296 counts/100ml); creek mouth (483
counts/100ml); and the Votua bridge with 751 counts/100ml. This
correlates to a 5-8 fold increase from the dam and downstream.
161
In terms of drinking water supply, less than 1 count/100ml of faecal
coliform and E/coli were observed for the Votua Housing tap water.
The Votua Village drinking tap water reached 58 counts/100ml of
faecal coliform and 48 counts/100ml of E/coli. A nearby diving
centre‘s (i.e. Mike‘s Divers) drinking tap water obtained 33 and 21
counts/100ml of faecal coliform and E/coli, respectively.
Conductivity increased from the dam to the creek mouth with a range
between 0.11mS and 0.23mS. In terms of mean stream flow it was
generally ‗fast‘ at the dam and upper housing; ‗medium‘ at lower
housing; and ‗slow‘ at the bridge and creek mouth. Total suspended
solids at the dam was 8.3 mg/L but elevated to the creek mouth with
14.3 mg/L. Votua Bridge displayed the highest suspended solids level
of 15.3 mg/L. Biological oxygen demand (BOD) showed no variance
between sites as all monitored locations yielded <18 mg/L.
Ammonia concentration ranged from 3.1µmol/L at the dam to
11.4µmol/L at the bridge. Nitrate also showed a similar trend to
ammonia at all the monitored sites along the Votua Creek with a low
of 1.96µmol/L at the dam to a high of 11.9µmol/L at the bridge.
Nitrite was least at the dam with 0.46µmol/L and increased
downstream to a maximum of 1.12µmol/L at the creek mouth. Total
kjeldahl nitrogen was least from the dam and relatively increased to
the creek mouth ranging from 7.2 to 23.8 µmol/L. Total inorganic
162
nitrogen reached a maximum at the bridge with 24.4µmol/L with the
lowest concentration of 5.5µmol/L at the dam. Phosphate
concentrations for the Votua Creek were 0.36µmol/L at the dam,
followed by upper housing with 0.39µmol/L, 0.78µmol/L for lower
housing, Votua bridge with 0.99µmol/L, and the creek mouth with
1.36µmol/L. Total phosphorus also highlighted similar trend with the
dam showing the lowest value of 0.44µmol/L. Upper housing had
0.50µmol/L; lower housing 0.67µmol/L; 0.93µmol/L for Votua bridge
whilst the creek mouth attained 1.35µmol/L.
Moreover, results for the greywater treatment drum experiment
signified that there would be little variation between one drum per
household (i.e. High Loading drum) and two drum per household (i.e.
Low Loading drum) in how they treat conductivity, dissolved oxygen,
total dissolved solid and E/coli, assuming that households dispose
only three large periodic doses per day into the onsite drum greywater
treatment systems. Also for monitoring period 1, results indicated that
one drum per household would be more efficient than two drum per
household in terms of faecal coliform, total suspended solids and total
kjeldahl nitrogen treatment. However for phosphorus and biological
oxygen demand treatment, the two drum per household would be
more efficient.
For monitoring period 2, which assumes that household greywater
discharges at Votua Village is achieved by moderate flushes every two
163
hours per day, then the two drum per household (i.e. low loading
drum) would be the ideal set up to improve salinity, dissolved oxygen
and pH as opposed to one drum per household (i.e. high loading). For
total suspended solids and total kjeldahl nitrogen treatment, the one
drum per household (i.e. high loading) would be the ideal system
choice. However, the choice of either one drum per household (high
loading) or two drum per household (low loading drum) in Votua does
not necessarily matter for elimination of other crucial water quality
parameters including faecal coliform, E/coli, phosphorus and
biological oxygen demand.
Data for monitoring period 3 showed that by assuming the greywater
discharge from Votua households to be emitted every forty minutes in
an average of 12 small doses per day; then the two drum per
household (i.e. low loading drum) set up would be more efficient in
improving dissolved oxygen, pH, faecal coliform, phosphorus and
biological oxygen demand whilst the one drum per household (i.e. high
loading drum) set up would enhance total kjeldahl nitrogen and E.coli
removal.
Furthermore, a gauge of the degree of soil-interface clogging test
indicated that for a standard high loading dosage volume of 85 litres
of artificial greywater solution, it took 73 minutes to elute in the clean
system and 118 minutes for the same volume at the end of the
experiment. The difference was 45 minutes. For a standard low
164
loading dosage volume of 42.5 litres of artificial greywater solution, it
took 38 minutes to elute for the clean system and lasted 66 minutes
for the clogged system at the end of the experiment. The difference
was 28 minutes.
Finally the wastewater treatment initiatives monitored under this
study including the greywater treatment drum experiment were the
first of their kind to be trialled in Fiji. Results obtained were difficult
to compare and contrast with local examples either in Fiji or in the
small Pacific Island Countries; except in countries of differing climate.
Therefore there is great need for further research into such
wastewater treatment initiatives such as the Tagaqe constructed
wetland and Crusoe‘s Resorts system as well as the greywater
treatment drums regardless of their potential efficiency, to ascertain a
better understanding of how they operate. Nevertheless the
constructed wetland and Crusoe‘s treatment system have a significant
potential to be promoted in other coastal communities and tourist
hotels in Fiji and elsewhere throughout the Pacific Islands.
165
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Appendix A: Tagaqe Constructed Wetland Monitoring
Date of Sampling: 15 June 2005 (Tagaqe wetland)
Influent Effluent % removal
Temperature (˚C) 24.6 23.7 -
Salinity (ppt) 0.0 0.0 -
Dissolved Oxygen (mg/L) 0.13 0.41 -
Conductivity (mS) 0.03 0.04 -
pH 6.78 6.82 -
Faecal Coliform (c/100ml) 9.29 x 106 6.85 x 104 99.3
TSS (mg/L) 1254.5 38.5 97.0
Mean BOD5 (mg/l) 452 18 96.0
Mean NH3-N (µM) 2564.3 662.1 74.2
Mean NO3-N (µM) 2.14 5.57 -
TIN (µM) 2566.44 667.67 74.0
Mean PO4-P (µM) 337.7 79.0 76.6
N : P ratio 8 8
Date of Sampling: 18 July 2005 (Tagaqe wetland)
Influent Effluent % removal
Temperature (˚C) 25.8 24.9 -
Salinity (ppt) 0.0 0.0 -
Dissolved Oxygen (mg/L) 0.16 0.47 -
Conductivity (mS) 0.03 0.05 -
pH 6.76 6.81 -
Faecal Coliform (c/100ml) 4.5 x 106 1.2 x 103 99.97
TSS (mg/L) 809 47 94.2
Mean BOD5 (mg/L) 265 18 93.2
Mean NH3-N (µM) 6242.86 814.29 87.0
Mean NO3-N (µM) 0.21 5.0 -
TIN (µM) 6243.07 819.29 86.9
TKN (µM) 10214.3 735.7 92.8
PO4-P (µM) 406.45 51.61 87.3
TP (µM) 598.1 57.4 90.4
N:P ratio 15 16
Date of Sampling: 20 October 2005 (Tagaqe wetland)
Influent Effluent % removal
Temp (˚C) 31 26.2 -
Salinity (ppt) 0.0 0.0 -
DO (mg/l) 0.25 0.85 -
Cond (mS) 0.01 0.002 -
pH 6.79 6.84 -
Mean FC (c/100ml) 2.6 x 106 1.9 x 104 99.3
Mean TSS (mg/l) 434 11 97.5
Mean BOD (mg/l) 182 18 90.1
Mean NH3-N (µM) 2314.29 470 79.7
Mean NO3-N (µM) 2.14 5.21 -
TIN (µM) 2316.43 475.21 79.5
TKN (µM) 7985.71 1907.14 76.1
Mean PO4-P (µM) 330 76.45 76.8
TP (µM) 445.16 141.94 68.1
N:P ratio 7 6
185
Date of Sampling: 11 May 2006 (Tagaqe wetland)
influent effluent % removal
Temp (˚C) 29.8 28.4 -
Sal (ppt) 0.0 0.0 -
DO (mg/l) 0.14 0.93 -
Cond (mS) 0.03 0.01 -
pH 6.81 6.85 -
FC (c/100ml) 2.3 x 105 4.8 x 103 97.9
BOD (mg/l) 500 18 96.4
NH3-N (µM) 4471.43 565.36 87.4
NO3-N (µM) 0.32 0.64 -
NO2-N (µM) 0.14 0.07 50.0
TIN (µM) 4471.89 566.07 87.3
PO4-P (µM) 359.35 71.94 80.0
N:P ratio 12 8
Date of Sampling: June 2006 (Tagaqe wetland)
influent Effluent % removal
Temp (ºC) 29.3 29.2 -
Salinity (ppt) 0.0 0.0 -
DO (mg/L) 0.19 0.89 -
Cond (mS) 0.04 0.01 -
pH 6.83 6.89 -
Mean FC (c/100ml) 9.0 x 106 4.8 x 104 99.5
Mean BOD (mg/L) 942 18 98.1
Mean TSS (mg/L) 855 22 97.4
Mean NH3-N (µM) 5738.57 740.7 87.1
Mean NO3-N (µM) 2.64 10.1 -
Mean NO2-N (µM) 0.14 0.07 50.0
TIN (µM) 5741.35 750.87 86.9
TKN (µM) 8272.4 2357.1 71.5
Mean PO4-P (µM) 282 92.97 67.0
TP (µM) 690.3 331.29 52.0
N:P ratio 8 8
Date of Sampling: 5 July 2006 (Tagaqe wetland) influent effluent % removal
Temp (ºC) 28.6 30.3 -
Sal (ppt) 0.0 0.0 -
DO (mg/l) 0.15 0.96 -
Cond (mS) 0.03 0.01 -
pH 6.79 6.86 -
FC (c/100ml) 1.6 x 105 3.5 x 104 78.1
E.coli (c/100ml) 9.0 x 105 3.5 x 104 96.1
BOD (mg/L) 211 18 91.5
TSS (mg/l) 420 10 97.6
NH3-N (µM) 4607.14 582.14 87.4
NO3-N (µM) 0.36 7.14 -
NO2-N (µM) 0.14 0.07 50.0
TIN (µM) 4607.64 589.35 87.2
TKN (µM) 6135.71 4064.29 33.8
PO4-P (µM) 301.29 70.32 76.7
TP (µM) 316.77 92.26 70.9
N:P ratio 15 6
186
Date of Sampling: September 2006 (Tagaqe wetland)
influent Effluent % removal
Temp (ºC) 30.1 27.5 -
Salinity (ppt) 0.0 0.0 -
DO (mg/L) 0.15 0.89 -
Cond (mS) 0.30 0.005 -
pH 6.84 6.87 -
Mean FC (c/100ml) 1.6 x 106 7.0 x 103 99.6
E.coli (c/100ml) 1.6 x 106 4.0 x 103 99.8
Mean BOD (mg/L) 28 18 30.8
Mean TSS (mg/L) 291 51 82.5
Mean NH3-N (µM) 6014.29 1164.29 80.6
Mean NO3-N (µM) 0.29 7.86 -
Mean NO2-N (µM) 0.14 0.07 50.0
Total Inorganic N (µM) 6014.72 1172.22 80.5
TKN (µM) 6121.43 1271.43 79.2
Mean PO4-P (µM) 525.81 110.97 78.9
TP (µM) 564.52 126.45 77.6
N:P ratio 11 11
Date of Sampling: 8 October 2006 (Tagaqe wetland)
influent Effluent % removal
Temp (˚C) 29.5 27.9 -
Salinity (ppt) 0.0 0.0 -
DO (mg/L) 0.13 1.14 -
Cond (mS) 0.45 0.01 -
pH 6.82 6.88 -
Mean FC (c/100ml) 1.6 x 106 7.0 x 103 99.6
Mean E.coli 1.6 x 106 7.0 x 103 99.6
Mean BOD (mg/L) 18 12 33
Mean TSS (mg/L) 2143 35 98.4
Mean NH3-N (µM) 4821.4 1385.7 71.3
Mean NO3-N (µM) 0.21 7.14 -
Mean NO2-N (µM) 0.14 0.07 50.0
Total Inorganic N (µM) 4821.75 1392.91 71.1
TKN (µM) 10285.71 1,650 84.0
Mean PO4-P (µM) 593.55 215.81 63.6
TP (µM) 619.35 250.32 59.6
N:P ratio 8 6
187
Appendix B: Crusoe’s Resort STP Monitoring Date of Sampling: 20 October 2005 (Crusoe’s STP)
Influent Effluent % removal
Temp (ºC) 32 31.4 -
Salinity (ppt) 0.0 0.0 -
DO (mg/l) 0.14 4.37 -
Cond (mS) 0.116 0.01 -
pH 6.76 6.85 -
Mean FC (c/100ml) 1.2 x 105 8.6 x 103 92.8
Mean TSS (mg/l) 16 3 81.3
Mean BOD (mg/l) 85 18 78.8
Mean NH3-N (µM) 3185.7 470 85.2
Mean NO3-N (µM) 55.7 15 73.1
TIN (µM) 3241.4 485 85.0
TKN (µM) 3557.14 1078.57 69.7
Mean PO4-P (µM) 180 31 82.8
TP (µM) 206.45 109.68 46.9
N:P ratio 18 16
Date of Sampling: June 2006 (Crusoe’s STP)
influent effluent % removal
Temp (ºC) 29.6 30.2 -
Sal (ppt) 0.0 0.0 -
DO (mg/l) 0.15 4.64 -
Cond (mS) 1.12 0.02 -
pH 6.77 6.96 -
FC (c/100ml) 9.0 x 105 7.0 x 104 92.2
TSS (mg/L) 18 4 77.8
BOD (mg/L) 87 18 79.3
NH3-N (µM) 3498.57 618.57 82.3
NO3-N (µM) 7.14 3.57 50.0
NO2-N (µM) 1.43 0.71 50.3
TIN (µM) 3507.14 622.85 82.2
PO4-P (µM) 166.45 90.65 45.5
N:P ratio 21 7
Date of Sampling: 5 July 2006 (Crusoe’s STP)
influent effluent % removal
Temp (ºC) 30.3 30.7 -
Sal (ppt) 0.0 0.0 -
DO (mg/l) 0.25 3.79 -
Cond (mS) 0.97 0.015 -
pH 6.76 6.87 -
FC (c/100ml) 7.4 x 105 3.9 x 104 94.7
E.coli (c/100ml) 6.6 x 105 3.8 x 104 94.2
BOD (mg/L) 76 18 76.3
TSS (mg/l) 26 5 80.8
NH3-N (µM) 3685.71 573.57 84.4
NO3-N (µM) 7.14 2.86 60.0
NO2-N (µM) 0.14 0.07 50.0
TIN (µM) 3692.99 576.5 84.4
PO4-P (µM) 223.2 68.71 69.2
N:P ratio 17 8
188
Date of Sampling: August 2006 (Crusoe’s STP)
influent Effluent % Removal
Temp (ºC) 29.8 31.2 -
Salinity (ppt) 0.0 0.0 -
DO (mg/l) 0.18 4.43 -
Cond (mS) 0.69 0.02 -
pH 6.70 6.92 -
FC (c/100ml) 2.4 x 106 9.0 x 104 96.3
E.coli (c/100ml) 2.4 x 106 9.0 x 104 96.3
Mean TSS (mg/L) 61 42 31.1
Mean BOD (mg/l) 25 18 28.0
NH3-N (µM) - - -
NO3-N (µM) - - -
NO2-N (µM) - - -
TIN (µM) - - -
TKN (µM) 1971.43 1235.71 37.3
PO4-P (µM) - - -
TP (µM) - - -
Date of Sampling: September 2006 (Crusoe’s STP)
influent Effluent % Removal
Temp (ºC) 30.0 30.2 -
Salinity (ppt) 0.0 0.0 -
DO (mg/l) 0.14 3.75 -
Cond (mS) 0.84 0.018 -
pH 6.82 6.89 -
FC (c/100ml) 2.8 x 105 8.0 x 104 71.4
E.coli (c/100ml) 5.0 x 104 3.0 x 104 40.0
Mean TSS (mg/L) 55 10 81.8
Mean BOD (mg/l) 45 18 60.0
NH3-N (µM) 2407.14 1828.57 24.0
NO3-N (µM) 7.14 2.86 60.0
NO2-N (µM) 0.14 0.07 50.0
TIN (µM) 2414.42 1831.5 24.1
TKN (µM) 2578.57 1728.57 33.0
PO4-P (µM) 311.94 68.06 78.2
TP (µM) 319.68 98.06 69.3
N:P ratio 8 27
189
Appendix C: Coral Coast Water Quality Monitoring
Date of Sampling: 18 July 2005 (Coral Coast Sites)
Site Place GPS location NO3-N (µM)
NH3-N (µM)
PO4-P (µM)
TIN:P ratio
1 Fijian Resort: ocean side 18-08.62S; 177-25.76E 6.43 0.97 0.49 15
2 Outrigger Resort: western side 18-10.82S; 177-33.08E 6.41 0.49 0.18 38
3 Tubakula Resort: eastern side 18-10.86S; 177-33.46E 5.53 0.61 0.37 17
5 East of Votua Village 18-12.69S; 177-42.89E 3.58 0.52 0.24 17
6 Tagaqe Village 18-11.91S; 177-39.75E 5.71 1.44 0.37 19
7 Sovi Bay beach 18-12.30S; 177-36.39E 5.23 0.85 0.33 18
8 Hideaway Resort: western side 18-11.92S; 177-39.32E 6.63 6.24 0.26 49
9 Front of Naviti Resort 18-12.31S; 177-41.84E 4.85 0.92 0.25 23
10 West of Komave Village 18-13.38S; 177-45.71E 2.48 0.98 0.31 11
11 Tabua Sands Resort 18-11.62S; 177-37.89E 5.73 0.62 0.28 23
12 Vatukarasa Bay 18-10.85S; 177-36.22E 4.95 0.65 0.45 12
13 Malevu Village: eastern side 18-10.85S; 177-33.62E 5.93 0.81 0.33 20
14 Crows Nest Resort 18-10.65S; 177-32.60E 6.57 0.84 0.35 21
15 Korotogo Bridge (River water) 18-10.67S; 177-32.59E 6.13 5.37 1.03 11
16 Matai Kandavu Beach 18-10.77S; 177-31.05E 2.99 0.57 0.10 36
17 Between Malevu/ Vatukarasa Villages 18-11.17S; 177-33.57E 4.86 0.49 0.31 17
18 Warwick Hotel 18-13.69S; 177-44.37E 11.50 0.99 0.54 23
mean 5.62 1.37 0.36 19
190
Date of Sampling: 20 October 2005 (Coral Coast Sites)
Site Place GPS location Time Temp (ºC)
Sal (ppt)
DO (mg/l)
Cond. (mS)
NO3-N (µM)
NH3-N (µM)
PO4-P (µM)
TIN:P ratio
5 Votua Village 18-12.69S 177-42.89E
1219 30 30.3 6.2 50.9 2.38 0.32 0.18 15
6 Tagaqe Village 18-11.91S 177-39.75E
1327 20.2 32.7 6.8 32.24 1.57 0.79 0.32 8
8 Hideaway Resort 18-11.92S 177-39.32E
1336 31.6 33.2 5.67 40 3.63 3.24 0.23 29
9 Naviti Resort 18-12.31S 177-41.84E
1230 32 30.3 6.23 54.2 4.43 0.87 0.25 21
10 Komave Village 18-13.38S 177-45.71E
1143 24 31.1 4.97 22.25 2.18 0.83 0.28 10
11 Tabua Sands 18-11.62S 177-37.89E
1351 23 31.6 6.12 48.3 3.41 0.48 0.31 11
18 Warwick Hotel 18-13.69S 177-44.37E
1202 32.3 29.4 6.7 53.5 4.5 0.49 0.33 15
mean 27.6 31.2 6.1 43.06 3.15 1.0 0.27 15
191
Date of Sampling: 11 May 2006 (Coral Coast Sites) Site Place Time Tide
Sal
(ppt)
Temp
(ºC)
DO
mg/l
Cond
(mS)
NO3-N
(µM)
NH3-N
(µM)
NO2-
N(µM)
PO4-
P(µM)
N:P
ratio
1 Fijian Resort [18-08.62S;177-25.76E]
1110 H+6 33.4 28.8 5.87 54.5 1.86 0.489 0.13 0.29 9
2 Outrigger [18-10.82S;177-33.08E] 1210 H-5 33 28.6 7.48 54.3 5.35 0.09 0.14 0.42 13
3 Tubakula [18-10.86S;177-33.46E] 1220 H-5 32.8 28.9 7.68 54.2 4.93 0.16 0.10 0.32 16
5 Votua Village [18-12.69S;177-
42.89E]
1410 H-3 31.3 28.9 5.67 56.2 1.97 0.22 0.10 0.20 11
6 Tagaqe Village [18-11.91S;177-
39.75E]
1321 H-4 32.6 30.8 6.81 52.3 1.71 1 0.32 0.23 13
7 Sovi Bay [18-12.30S;177-36.39E] 1246 H-5 34.4 29.1 6.6 56.7 3.96 0.10 0.06 0.27 15
8 Hideaway Resort
18-11.92S;177-39.32E
1315 H-4 33.2 28.9 7.51 43.5 3.23 0.28 0.12 0.24 15
9 Naviti Resort [18-12.31S;177-
41.84E]
1343 H-4 34.6 32.4 6.64 53.8 3.76 0.31 0.08 0.26 16
10 Komave Village 18-13.38S;177-45.71E
1430 H-3 32 27.7 5.75 51.2 2.11 0.23 0.11 0.29 8
11 Tabua Sands [18-11.62S;177-
37.89E]
1305 H-4 31 29.6 6.57 54.1 2.62 0.27 0.12 0.33 9
12 Vatukarasa 18-10.85S;177-36.22E 1253 H-5 31.7 28.4 7.77 48.9 1.72 0.13 0.09 0.17 11
13 Malevu Village
18-10.85S;177-33.62E
1230 H-5 34.7 28.9 6.34 56.8 1.20 0.20 0.10 0.15 10
14 Crows Nest [18-10.65S;177-
32.60E]
1205 H-5 33.6 28.9 7.68 55.2 1.91 0.33 0.08 0.19 12
15 Korotogo Bridge
18-10.67S;177-32.59E
1154 H-6 27 28.4 3.50 47.96 6.79 4.58 0.23 1.24 9
16 Matai Kadavu Beach 18-10.77S;177-31.05E
1130 H-6 31.2 30.6 6.32 26.14 0.60 0.12 0.07 0.14 6
17 Between Malevu & Vatukarasa
18-11.17S;177-33.57E
1240 H-5 29.7 30.3 6.78 43.45 1.09 0.16 0.09 0.20 7
18 Warwick Hotel
18-13.69S;177-44.37E
1355 H-4 33.1 29.8 6.64 43.9 2.12 0.29 0.08 0.31 8
mean 32.3 29.4 6.57 47.6 2.76 0.53 0.12 0.31 11
192
Date of Sampling: June 2006 (Coral Coast Sites) Site Place Time Tide Sal
(ppt)
Temp
(ºC)
DO
mg/l
Con
(mS)
NO3-N
(µM)
NH3-N
(µM)
NO2-N
(µM)
PO4-P
(µM)
T:P
1 Fijian Resort [18-08.62S;177- 25.76E]
1110 H+1 32.6 29.6 5.67 52.4 0.98 3.67 0.63 0.16 32
2 Outrigger Resort [18-10.82S;177-
33.08E]
1235 H+3 31.4 27.9 6.78 51.5 8.29 4.26 0.64 0.47 28
3 Tubakula Resort [18-10.86S;177-
33.46E]
1243 H+3 32.5 29.2 6.43 49.7 8.07 2.98 0.63 0.31 37
5 Votua Village [18-12.69S;177-
42.89E]
1408 H+4 32.8 27.5 5.88 54.3 3.89 6.48 0.64 1.08 10
6 Tagaqe Village [18-11.91S;177-
39.75E]
1321 H+3 32.1 30.3 5.92 46.7 1.86 1.14 0.30 0.71 5
7 Sovi Bay [18-12.30S;177-36.39E] 1258 H+3 33.3 31.2 5.75 55.4 3.06 4.27 0.63 0.38 21
8 Hideaway Resort [18-11.92S;177-39.32E]
1332 H+3 31.2 29.5 5.74 53.6 3.26 5.29 0.64 0.23 39
9 Naviti Resort[18-12.31S;177-
41.84E]
1355 H+4 30.7 31.5 5.69 52.1 2.56 9.14 0.64 0.58 21
10 Komave [18-13.38S;177-45.71E] 1430 H+4 31.6 29.8 5.30 53.4 2.78 4.31 0.58 0.46 16
11 Tabua Sands [18-11.62S;177-
37.89E]
1323 H+3 32.7 27.9 5.54 52.8 5.84 1.75 0.65 0.36 22
12 Vatukarasa [18-10.85S;177-36.22E] 1310 H+3 29.5 29.3 5.36 49.8 3.47 5.77 0.66 1.22 8
13 Malevu [18-10.85S;177-33.62E] 1250 H+3 33.4 29.6 6.65 55.2 13.5 3.18 0.64 0.83 20
14 Crows Nest [18-10.65S;177-32.60E] 1228 H+3 32.1 29.9 5.94 51.3 7.86 2.49 0.64 0.75 14
15 Korotogo Bridge [18-10.67S;177-
32.59E]
1222 H+3 27.6 29.3 3.43 43.8 4.11 7.50 0.70 1.14 11
16 Matai Kadavu Beach
[18-10.77S;177-31.05E]
1200 H+2 32.4 29.5 5.30 47.3 5.31 4.39 1.18 0.82 13
17 Between Malevu & Vatukarasa 18-11.17S;177-33.57E
1240 H+3 31.5 29.4 5.81 49.9 4.64 3.29 0.39 0.53 15
18 Warwick Hotel
[18-13.69S;177-44.37E]
1355 H+4 32.9 30.3 6.13 53.6 0.97 3.33 0.63 0.32 15
mean 31.7 29.5 5.73 51.3 4.73 4.31 0.64 0.61 19
193
Date of Sampling: 5 July 2006 (Coral Coast Sites)
Site Place Time Tide Sal (ppt)
Temp. (ºC)
DO (mg/l)
Cond. (mS)
NO3-N (µM)
NH3-N (µM)
NO2-N (µM)
PO4-P (µM)
TIN:P ratio
5 Votua Village 18-12.69S 177-42.89E
1210 H-2 31.9 30.7 5.85 56.1 2.45 4.23 0.21 0.83 8
6 Tagaqe Village
18-11.91S 177-39.75E
1355 H+0 32.9 31.5 5.76 52.4 2.07 1.0 0.23 0.64 5
7 Sovi Bay 18-12.30S 177-36.39E
1445 H +1 32.5 29.9 5.23 49.8 2.71 3.34 0.31 0.41 10
8 Hideaway Resort 18-11.92S 177-39.32E
1410 H+0 34.5 30.6 6.41 51.5 3.13 3.71 0.29 0.43 15
9 Naviti Resort 18-12.31S 177-41.84E
1330 H-1 30.3 32.3 5.32 47.2 2.33 4.11 0.15 0.51 13
10 Komave Village 18-13.38S 177-45.71E
1120 H-3 31.7 27.4 6.11 53.7 2.08 3.65 0.19 0.42 14
11 Tabua Sands 18-11.62S 177-37.89E
1420 H+0 33.1 28.1 5.57 51.3 1.92 1.78 0.14 0.36 11
12 Vatukarasa 18-10.85S 177-36.22E
1435 H+1 33.8 31.3 6.72 53.4 2.43 3.22 0.26 0.47 13
18 Warwick Hotel 18-13.69S 177-44.37E
1140 H-2 31.4 30.6 6.23 53.2 1.47 3.54 0.12 0.33 15
mean 32.5 30.3 5.91 52.1 2.29 3.18 0.21 0.49 12
194
Appendix D: Votua Creek Water Quality Monitoring
Monthly Sampling: June 2006 (Votua Creek) Site Faecal Coliform (c/100ml) NH3-N (µM) NO3-N (µM) NO2-N (µM) PO4-P (µM)
Votua dam 168 - - - -
Upper housing 273 5.34 12.50 0.71 0.41
Lower housing 1100 14.0 12.57 0.75 1.20
Votua bridge 300 14.14 13.36 0.95 1.17
Votua creek mouth 1700 10.07 9.14 1.0 1.42
Monthly Sampling: July 2006 (Votua Creek)
Site FC c/100ml
E.coli c/100ml
Sal ppt
Temp ºC
Cond mS
pH TSS mg/l
Flow Tide BOD mg/l
NH3 (µM)
NO3 (µM)
NO2 (µM)
TIN (µM)
TKN (µM)
PO4 (µM)
TP (µM)
Votua dam
34 34 0.1 21.7 0.108 7.77 4 Fast flow
rising <18 3.32 2.34 0.37 6.03 10 0.33 0.45
Upper housing
125
125 0.1 22.4 0.110 7.42 2 medium rising <18 3.81
5.98
0.47
10.26 12 0.37
0.67
Lower housing
320
229 0.1 22.7 0.112 7.34 7 medium rising <18 12.79
13.37
0.71
26.87 28.4 0.61
0.65
Votua bridge
54
54 0.2 23.4 0.121 7.4 6 slow rising <18 13.0
12.52
0.95
26.47 27.7 0.92
0.97
Votua creek mouth
300
200 15.6 24.2 0.235 7.8 6 slow rising <18 10.53
9.47
0.99
20.99 24.1 1.26
1.29
Housing tap water
<1 <1
village tap water
60 30
Mike diver‘s tap water
29 23
Housing dug out
well
57 21
195
Monthly Sampling: August 2006 (Votua Creek) Site FC
c/100ml E.coli c/100ml
Sal ppt
Temp ºC
Cond mS
pH TSS mg/l
Flow Tide BOD mg/l
NH3 (µM)
NO3 (µM)
NO2 (µM)
TIN (µM)
PO4 (µM)
TP (µM)
Votua dam 57 50 0.0 18.4 0.111 7.71 2 Fast Rising <18 2.73 1.67 0.46 4.86 0.40 0.40
Upper housing 60 60 0.0 16.9 0.107 7.22 3 Fast Rising <18 4.51 3.89 0.49 8.89 0.40 0.42
Lower housing 300 240 0.0 21.2 0.113 7.24 5 Fast Rising <18 7.90 8.91 0.73 17.54 0.64 0.65
Votua bridge 800 800 0.1 20.7 0.125 7.10 7 Slow Rising <18 8.79 10.74 1.11 20.64 0.89 0.88
Votua creek mouth
500 500 1.7 26.1 0.224 7.78 6 slow rising <18 9.48 11.81 1.22 22.51 1.30 1.31
Housing tap water
2 <1
Votua village tap water
36 36
Mike diver‘s tap water
37 19
Monthly Sampling: September 2006 (Votua Creek)
FC c/100ml
E.coli c/100ml
Sal ppt
Temp ºC
Cond mS
pH TSS mg/l
Flow Tide BOD mg/l
NH3
(µM) NO3
(µM) NO2
(µM) TIN
(µM) TKN
(µM) PO4
(µM) TP
(µM) Votua dam
300 200 0.0 17.7 0.112 7.7 19 Fast Rising <18 3.29 1.88 0.56 5.73 4.35 0.34 0.46
Upper
housing
200 220 0.0 15.1 0.117 7.5 24 Fast Rising <18 5.14 3.41 0.56 9.11 11.2 0.38 0.41
Lower housing
420 420 0.0 18.3 0.121 7.34 30 Medium Rising <18 8.16 9.29 0.80 18.25 24.55 0.66 0.71
Votua bridge
1.5 x
103
1.4 x 103
0.0 18.6 0.127 7.30 33 Fast Rising <18 9.62 10.98 1.27 21.87 18.90 0.97 0.93
Votua
creek mouth
750 750 1.0 17.4 0.231 7.87 31 Medium rising <18 11.06 12.21 1.28 24.55 23.41 1.46 1.44
Housing
tap water
<1 <1
Votua
village tap water
78 78