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Domestic Wastewater Treatment in
Developing Countries
Domestic Wastewater Treatment in
Developing Countries
Duncan Mara
London • Sterling, VA
First published by Earthscan in the UK and USA in 2004
Copyright © Duncan Mara, 2003
All rights reserved
ISBN: 1-84407-019-0 paperback
1-84407-020-4 hardback
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A catalogue record for this book is available from the British Library
Library of Congress Cataloging-in-Publication Data
Mara, D. Duncan (David Duncan), 1944Domestic wastewater treatment in developing countries / Duncan Mara.
p. cm.
Includes bibliographical references and index.
ISBN 1-84407-020-4 (alk. paper) – ISBN 1-84407-019-0 (pbk. : alk. paper)
1. Sewage disposal–Developing countries. 2. Sewage--Purification–Developing
countries. I. Title.
TD627.M37 2004
This book is printed on elemental chlorine free paper
List of Figures and Tables
Principal Notation
List of Acronyms and Abbreviations
What is Domestic Wastewater and Why Treat It?
Origin and composition of domestic wastewater
Characterization of domestic wastewater
Wastewater collection
Why treat wastewater?
Investment in wastewater treatment
Excreta-related Diseases
Environmental classification of excreta-related diseases
Global burden of excreta-related diseases
Essential Microbiology and Biology
Bacteria and Archaea
Freshwater micro-invertebrates
Effluent Quality
Wastewater treatment objectives
Wastewater re-use
Discharge to inland waters
Discharge to coastal waters
BOD Removal Kinetics
First-order kinetics
Hydraulic flow regimes
Limitations of simple first-order kinetics
Worked examples
vi Domestic Wastewater Treatment in Developing Countries
Domestic Wastewater Treatment Options
Sustainability issues
Appropriate wastewater treatment options
Sustainable wastewater treatment options
Domestic Wastewater Flows and Loads
Domestic wastewater flows
Domestic wastewater loads
Future projections
Preliminary Treatment
Grit removal
Flow measurement
Waste Stabilization Ponds
Types and functions of WSP
Advantages of WSP
Perceived disadvantages of WSP
WSP usage
High altitude WSP
WSP or other treatment processes?
Macrophyte ponds
Advanced pond systems
Anaerobic Ponds
High-rate anaerobic ponds
Anaerobic ponds in series
Design example
Facultative Ponds
Algal biomass
Purple ponds
Wind-powered pond mixers
Design examples
Maturation Ponds
Pathogen removal mechanisms
Design for E coli removal
Design for helminth egg removal
Contents vii
BOD removal
Nutrient removal
Pond effluent polishing
Design example
Physical Design of WSP
Pond location
Geotechnical considerations
Pond lining
Pond geometry
Inlet and outlet structures
By-pass pipework
Anaerobic pond covers
Operator facilities
Upgrading and extending existing WSP
Operation and Maintenance of WSP
Start-up procedures
Routine maintenance
Desludging and sludge disposal
Staffing levels
Pond rehabilitation
Monitoring and Evaluation of WSP
Effluent quality monitoring
Evaluation of pond performance
Data storage and analysis
Wastewater Storage and Treatment Reservoirs
Single reservoirs
Sequential batch-fed reservoirs
Hybrid WSP–WSTR system
Design examples
Constructed Wetlands
Subsurface-flow wetlands
Wetlands or waste stabilization ponds?
Upflow Anaerobic Sludge Blanket Reactors
Treatment principles
UASBs or anaerobic ponds?
viii Domestic Wastewater Treatment in Developing Countries
Fly control
Design example
Simple Activated Sludge Variants
Aerated lagoons
Oxidation ditches
Wastewater Re-use in Agriculture
Why re-use wastewater?
Public health protection
Crop health
Treatment options for re-use
Quantitative microbial risk analysis
Irrigation with untreated wastewater
Wastewater Re-use in Aquaculture
What is aquaculture?
Wastewater-fed aquaculture
Wastewater-fed fishpond design
Integrated agricultural–aquacultural re-use
Design example
List of Figures and Tables
Composition of Domestic Wastewater
Four-year old African girl with a Distended Abdomen
The Tree of Life
Common Bacterial Shapes
The Bacterial Batch-culture Growth Curve
The Catabolic, Anabolic and Autolytic Reactions of Aerobic
Microbiological Oxidation
3.5 Five of the Commonest Ciliated Protozoa in Wastewater
Treatment Works
3.6 Micro-invertebrates Used to Assess the Biological Quality of
Tropical Waters
4.1 The Dissolved Oxygen Sag Curve
4.2 Discharge of an Effluent into a River
5.1 Generalized BOD Curves
5.2 Thirumurthi Chart for the Wehner–Wilhelm Equation
5.3 Typical Tracer Study Results
7.1 Diurnal Variation of Wastewater Flow and Load at Nakuru,
8.1 Simple Manually Raked Screen
8.2 Mechanically Raked Screen
8.3 Flow Elements in a Parabolic Channel
8.4 Trapezoidal Approximation to a Parabolic Section
9.1 One of the Phase II 21-ha Primary Facultative Ponds at Dandora,
Nairobi, Kenya
9.2 Algal–bacterial Mutualism in Facultative and Maturation Ponds
9.3 Typical WSP Layout
9.4 Variation of Discount Rate and Land Price below which WSP
are the Cheapest Treatment Option
9.5 The Phase I WSP at Dandora, Nairobi, Kenya
9.6 The ‘55 East’ WSP Series at Werribee, Melbourne, Australia
9.7 The Mangere WSP, Auckland, New Zealand, in 1996
10.1 Anaerobic Pond, with Partial Scum Coverage, at Ginebra,
Valle del Cauca, Southwest Colombia
10.2 Variation of the Proportions of Hydrogen Sulphide, Bisulphide
and Sulphide with pH in Aqueous Solutions
x Domestic Wastewater Treatment in Developing Countries
10.3 High-rate Anaerobic Pond with a Mixing Pit
11.1 Diurnal Variation of Dissolved Oxygen in a Facultative Pond
11.2 Variation of Surface BOD Loading on Facultative Ponds with
Temperature According to Equations 11.2 and 11.3
11.3 Diurnal Variation in Facultative Pond Effluent Quality
11.4 Variation of Chlorophyll a with Surface BOD Loading on
Primary Facultative Ponds in Northeast Brazil
11.5 Photosynthetic Purple Sulphur Bacteria
12.1 Variation of kB with Surface BOD Loading on Primary
Facultative Ponds in Northeast Brazil
12.2 Variation of kB with In-pond Chlorophyll a Concentration in
Primary Facultative Ponds in Northeast Brazil
13.1 Embankment Protection by Concrete Cast in situ
13.2 Embankment Protection by Precast Concrete Slabs
13.3 Embankment Protection by Stone Rip-rap
13.4 Anaerobic Pond Lined with an Impermeable Plastic Membrane
13.5 Anchoring the Pond Liner at the Top of the Embankment
13.6 Calculation of Top and Bottom Pond Dimensions
13.7 Inlet Structure for Anaerobic and Primary Facultative Ponds
13.8 Inlet Structure on a Facultative Pond with Integral Scum Box
13.9 Inlet Structure for Secondary Facultative and Maturation Ponds
13.10 Outlet Weir Structure
13.11 By-pass Pipework for Anaerobic Ponds
13.12 Covered Anaerobic Pond at the Western Treatment Plant,
Melbourne, Australia
13.13 Partial View of the Al Samra WSP, Amman, Jordan
13.14 Fence and Warning Notice in English and Kiswahili at a Pond
Site in Nairobi, Kenya
13.15 Upgrading a WSP Series to Treat Twice the Original Flow
14.1 Sludge Depth Measurement by the ‘White Towel’ Test
14.2 Pond Desludging in Northern France
14.3 A Very Badly Neglected Facultative Pond in Eastern Africa
15.1 Details of Pond Column Sampler
16.1 Single WSTR in Israel
16.2 Wastewater Storage and Treatment Reservoir Systems
16.3 Sequential Batch-fed WSTR at Arad, Israel
17.1 A 100-m Long Subsurface-flow Constructed Wetland in Egypt
17.2 A Horizontal-flow Constructed Wetland at a Hotel in Kandy,
Sri Lanka
18.1 A UASB at Ginebra, Valle del Cauca, Southwest Colombia
18.2 Schematic Diagram of a UASB
18.3 Influent Distribution Channel and Distribution Boxes
18.4 Details of a Submerged Phase Separator
19.1 Sectional Perspective View of a Circular Biofilter
19.2 Distribution of Settled Wastewater on to a Rectangular Biofilter
List of Figures and Tables xi
19.3 Rectangular Biofilters with High-density Polyethylene Netting to
Control Fly Nuisance
20.1 An Aerated Lagoon
20.2 Floating ‘Aire-O2 Triton’ Aerator–mixer
20.3 Typical Oxidation Ditch Installation
21.1 Excess Prevalence of Ascaris and Hookworm Infections in
Sewage Farm Workers in India
21.2 Excess Intensity of Ascaris and Hookworm Infections in Sewage
Farm Workers in India
21.3 Ascaris Prevalence among Residents of Western Jerusalem,
21.4 Ascaris Prevalence among Residents of Selected German Cities
Immediately After the Second World War
21.5 Generalized Model Showing the Levels of Relative Risk to
Human Health Associated with Different Combinations of
Control Methods for the Use of Wastewater in Agriculture and
21.6 Drip Irrigation of Cotton with Maturation Pond Effluent at
Nicosia, Cyprus
21.7 Classification of Irrigation Waters Based on Conductivity and
Sodium Absorption Ratio
22.1 Some of the Kolkata East Wastewater-fed Fishponds
22.2 Harvesting Indian Major Carp from the Kolkata East
Wastewater-fed Fishponds
Composition of Human Faeces and Urine
Wastewater Strength in Terms of BOD5 and COD
Average BOD5 Contributions per Person per Day
Environmental Classification of Excreta-related Diseases
Major Excreta-related Pathogens Identified Since 1973
Global Diarrhoeal Disease and Geohelminthiases Statistics for
Micro-invertebrate Groups Used to Assess the Biological Quality
of Tropical Waters
Simplified Biotic Index for Tropical Waters
Normalized Unit Values for Dissolved Oxygen, Total Dissolved
Salt and Turbidity Used to Calculate WQImin
The UK Royal Commission’s Classification of River Water
The UK Royal Commission’s Standards for Wastewater Effluents
Discharged into Rivers
Effluent Quality Requirements for Domestic Wastewaters
Discharged into the Marine Environment of the Wider Caribbean
xii Domestic Wastewater Treatment in Developing Countries
BOD Removal Results in Primary Facultative Ponds in Northeast
Comparison of Factors of Importance in Wastewater Treatment
in Industrialized and Developing Countries
Costs and Land Area Requirements for WSP and Other
Treatment Processes
Excreted Pathogen Removals in WSP and Conventional
Treatment Processes
Design Values of Volumetric BOD Loadings on and Percentage
BOD Removals in Anaerobic Ponds at Various Temperatures
Variation of BOD Removal with BOD Loading and Retention
Time in Anaerobic Ponds in Northeast Brazil at 25ºC
Examples of Algal Genera Found in Facultative and Maturation
Performance of a Series of Five WSP in Northeast Brazil
Bacterial and Viral Removals in a Series of Five WSP in
Northeast Brazil
Settling Velocities for Parasite Eggs and Cysts
Helminth Egg Removal in Waste Stabilization Ponds in
Northeast Brazil
Reported Values of kB(20) and φ for Use in Equation 12.2
Performance Data for WSP with Different Depths and Length-toBreadth Ratios in Northeast Brazil at 25ºC
Recommended Staffing Levels for WSP Systems
Parameters to be Determined for Level 2 Pond Effluent Quality
Parameters to be Determined for the Minimum Evaluation of
WSP Performance
Operational Strategy for Three Sequential Batch-fed WSTR for
an Irrigation Season of Six Months
Solubility of Oxygen in Distilled Water at Sea Level at Various
Design Criteria for Oxidation Ditches in India and Europe
Crop Yields for Wastewater and Freshwater Irrigation in India
Recommended Maximum Concentrations of Boron in Irrigation
Waters According to Crop Tolerance
Recommended Maximum Metal Concentrations in Irrigation
Physicochemical Quality of Three Waste Stabilization Pond
Effluents in Israel
Values of N50 and α for Excreted Viral and Bacterial Pathogens
Percentage of Free Ammonia (NH3) in Aqueous Ammonia
(NH3 + NH4) Solutions at 1–25 °C and pH 7.0–8.5
This book is primarily written for final year undergraduate civil engineering
students in developing country universities, for post-graduate masters students
in environmental, public health and sanitary engineering, and for practising
engineers working in developing countries or working on wastewater
treatment projects in these countries. The primary emphasis of the book is on
low-cost, high-performance, sustainable domestic wastewater treatment
systems. Most of the systems described are ‘natural’ systems – so called because
they do not require any electromechanical power input. The secondary
emphasis is on wastewater re-use in agriculture and aquaculture – after all, it
is better to use the treated wastewater productively and therefore profitably,
rather than simply discharge it into a river and thus waste its water and its
nutrients. The human health aspects of wastewater use are obviously
important and these are covered in detail, including an introduction to
quantitative microbial risk analysis.
Over the last 30 or so years that I have been working on wastewater
engineering in developing countries, I have been helped by many colleagues
and friends. I particularly wish to express my gratitude to all of the following:
Professor Richard Feachem (University of California San Francisco and
Berkeley), Dr Mike McGarry (Cowater International, Ottawa), Emeritus
Professor Gerrit Marais (University of Cape Town), Professor Howard Pearson
(Universidade Federal do Rio Grande do Norte), Emeritus Professor Hillel
Shuval (Hebrew University of Jerusalem), Professor Sandy Cairncross and Dr
Ursula Blumenthal (London School of Hygiene and Tropical Medicine),
Emeritus Professor Takashi Asano (University of California Davis), Professor
Marcos von Sperling (Universidade Federal de Minas Gerais), Professor Peter
Edwards (Asian Institute of Technology) and Dr Andy Shilton (Massey
University); and at the University of Leeds: Emeritus Professor Tony Cusens,
Emeritus Professor Donald Lee, Professor Ed Stentiford, Dr Nigel Horan and
Dr Andy Sleigh. Advice on the content of Figure 3.1 was generously provided
by Dr Ian Head (University of Newcastle).
Docendo dedici. Many of my former doctoral students have made major
contributions, including Dr Rachel Ayres, Dr Harin Corea, Dr Tom Curtis, Dr
Martin Gambrill, Dr Steve Mills, Dr John Oragui, Dr Miguel Peña Varón,
Professor Salomão Silva, Dr David Smallman, Dr Rebecca Stott and Dr Huw
xiv Domestic Wastewater Treatment in Developing Countries
Finally, but most importantly, I wish to express a lifelong gratitude to
Kevin Newman, Emeritus Professor of Classics at the University of Illinois,
who taught me as a teenager how to think – the greatest gift a teacher can
Principal Notation
depth; dissolved oxygen deficit
number of helminth eggs
net evaporation
soluble BOD
first-order rate constant for BOD removal
first-order rate constant for surface reaeration
first-order rate for E coli removal
BOD; length
infectivity constant
volume; velocity
cell concentration
yield coefficient
oxygen consumed
coefficient of retardation; infectivity constant; ratio of oxygen
transfer in wastewater and tap water
ratio of oxygen solubility in wastewater and distilled water
sludge loading factor
dispersion number
retention time
first-order rate constant for soluble BOD removal
loading rate
specific growth rate
Arrhenius constant
xvi Domestic Wastewater Treatment in Developing Countries
maturation, mean, mixture
List of Acronyms and Abbreviations
advanced integrated pond system
advanced integrated wastewater ponding system
best available technology not entailing excessive cost
biochemical oxygen demand
cheapest available technology narrowly avoiding prosecution
carbonaceous BOD
chemical oxygen demand
dissolved oxygen
European Union
Food and Agriculture Organization
faecal coliforms
gross algal oxygen production
gross dissolved oxygen production
high-rate algal ponds
National River Conservation Directorate
operation and maintenance
photosynthetically active radiation
photosynthetic photon flux density
quantitative microbial risk analysis
solids retention time
suspended solids
theoretical oxygen demand
upflow anaerobic sludge blanket reactor(s)
United States Agency for International Development
United States Environmental Protection Agency
World Health Organization
waste stabilization pond(s)
wastewater storage and treatment reservoir(s)
What is Domestic Wastewater and
Why Treat It?
Domestic wastewater is the water that has been used by a community and
which contains all the materials added to the water during its use. It is thus
composed of human body wastes (faeces and urine) together with the water
used for flushing toilets, and sullage, which is the wastewater resulting from
personal washing, laundry, food preparation and the cleaning of kitchen
Fresh wastewater is a grey turbid liquid that has an earthy but inoffensive
odour. It contains large floating and suspended solids (such as faeces, rags,
plastic containers, maize cobs), smaller suspended solids (such as partially
disintegrated faeces, paper, vegetable peel) and very small solids in colloidal (ie
non-settleable) suspension, as well as pollutants in true solution. It is
objectionable in appearance and hazardous in content, mainly because of the
number of disease-causing (‘pathogenic’) organisms it contains (Chapter 2). In
warm climates wastewater can soon lose its content of dissolved oxygen and
so become ‘stale’ or ‘septic’. Septic wastewater has an offensive odour, usually
of hydrogen sulphide.
The composition of human faeces and urine is given in Table 1.1, and for
wastewater, in simpler form, in Figure 1.1. The organic fraction of both is
composed principally of proteins, carbohydrates and fats. These compounds,
particularly the first two, form an excellent diet for bacteria, the microscopic
organisms whose voracious appetite for food is exploited by public health
engineers in the microbiological treatment of wastewater. In addition to these
chemical compounds, faeces and, to a lesser extent, urine contain many
millions of intestinal bacteria and smaller numbers of other organisms. The
majority of these are harmless – indeed some are beneficial – but an important
minority is able to cause human disease (Chapter 2).
Sullage contributes a wide variety of chemicals: detergents, soaps, fats and
greases of various kinds, pesticides, anything in fact that goes down the kitchen
sink, and this may include such diverse items as sour milk, vegetable peelings,
tea leaves, soil particles (arising from the preparation of vegetables) and sand
2 Domestic Wastewater Treatment in Developing Countries
Table 1.1 Composition of Human Faeces and Urine
Quantity (wet) per person per day
Quantity (dry solids) per person per day
Approximate composition (%)
Organic matter
Phosphorus (as P2O5)
Potassium (as K2O)
Calcium (as CaO)
135–270 g
35–70 g
1.0–1.3 kg
50–70 g
Source: Gotaas (1956)
(used to clean cooking utensils). The number of different chemicals that are
found in domestic wastewater is so vast that, even if it were possible, it would
be meaningless to list them all. For this reason wastewater treatment engineers
use special parameters to characterize wastewaters.
As is explained more fully in Chapter 5, wastewaters are usually treated by
supplying them with oxygen so that bacteria can utilize the wastewater
contents as food. The general equation is:
wastewater + oxygen treated wastewater + new bacteria
The nature of domestic wastewater is so complex that it precludes its complete
analysis. However, since it is comparatively easy to measure the amount of
oxygen used by the bacteria as they oxidize the wastewater, the concentration
of organic matter in the wastewater can easily be expressed in terms of the
amount of oxygen required for its oxidation. Thus, if, for example, half a gram
of oxygen is consumed in the oxidation of each litre of a particular wastewater,
then we say that this wastewater has an ‘oxygen demand’ of 500 mg/l, by
which we mean that the concentration of organic matter in a litre of the
wastewater is such that its oxidation requires 500 mg of oxygen. There are
basically three ways of expressing the oxygen demand of a waste:
Theoretical oxygen demand (ThOD) – this is the theoretical amount of
oxygen required to oxidize the organic fraction of the wastewater
completely to carbon dioxide and water. The equation for the total
oxidation of, say, glucose is:
What is Domestic Wastewater and Why Treat It? 3
Source: Tebbutt (1998)
Figure 1.1 Composition of Domestic Wastewater
C6H12O6 + 6O2 6CO2 + 6H2O
With C = 12, H = 1 and O = 16, C6H12O6 is 180 and 6O2 is 192; we can thus
calculate that the ThOD of, for example, a 300 mg/l solution of glucose is
(192/180) x 300 = 321 mg/l. Because wastewater is so complex in nature its
ThOD cannot be calculated, but in practice it is approximated by the chemical
oxygen demand.
Chemical oxygen demand (COD) – this is obtained by oxidizing the
wastewater with a boiling acid dichromate solution. This process oxidizes
almost all organic compounds to carbon dioxide and water, the reaction
usually proceeding to more than 95 per cent completion. The advantage of
COD measurements is that they are obtained very quickly (within 3 hours),
but they have the disadvantages that they do not give any information on
the proportion of the wastewater that can be oxidized by bacteria, nor on
the rate at which bio-oxidation occurs.
Biochemical oxygen demand (BOD) – this is the amount of oxygen
required for the oxidation of a wastewater by bacteria. It is therefore a
measure of the concentration of organic matter in a waste that can be
oxidized by bacteria (‘bio-oxidized’ or ‘biodegraded’). BOD is usually
expressed on a 5-day, 20°C basis – that is as the amount of oxygen
consumed during oxidation of the wastewater for 5 days at 20°C. This is
because the 5-day BOD (usually written ‘BOD5’) is more easily measured
4 Domestic Wastewater Treatment in Developing Countries
than is the ultimate BOD (BODu), which is the oxygen required for the
complete bio-oxidation of the waste. (The reason for the seemingly
arbitrary choice of 20°C and 5 days for the measurement of BOD is given
in Chapter 4; see also Baird and Smith, 2002.) The correct concept of BOD
is fundamental to wastewater treatment, and a rigorous treatment of BOD
removal kinetics is given in Chapter 5.
From the foregoing it is apparent that:
ThOD > COD > BODu > BOD5
There is no general relationship between these various oxygen demands.
However, for untreated domestic wastewater a large number of measurements
have indicated the following approximate ratios:
BOD5/COD = 0.5
BODu/BOD5 = 1.5
The presence of industrial or agricultural wastewaters alters these ratios
Wastewater strength
The higher the concentration of organic matter in a wastewater, the ‘stronger’
it is said to be. Wastewater strength is often judged by its BOD5 or COD (Table
1.2). The strength of the wastewater from a community is governed to a very
large degree by its water consumption. Thus, in the US where water
consumption is high (350–400 l/person day) the wastewater is weak (BOD5 =
200–250 mg/l), whereas in tropical countries the wastewater is strong (BOD5
= 300–700 mg/l) as the water consumption is typically much lower (40–100
l/person day).
The other factor determining the strength of domestic wastewater is the
BOD (= amount of organic waste) produced per person per day. This varies
from country to country and the differences are largely due to differences in
the quantity and quality of sullage rather than of body wastes, although
variations in diet are important. A good value to use in developing countries is
40 g BOD5 per person per day (Table 1.3). In Brazil the BOD contribution per
person per day was found to vary with income – poor people produce less
Table 1.2 Wastewater Strength in Terms of BOD5 and COD
Very strong
BOD5 (mg/l)
COD (mg/l)
What is Domestic Wastewater and Why Treat It? 5
Table 1.3 Average BOD5 Contributions per Person per Day
Personal washing
Garbage disposala
Toilet – faeces
Total (average adult contribution)
Developing countries
Sources: Ligman et al (1974), Mara (1976)
a Sink-installed garbage grinder
b Includes allowance for food scraps
c Cleansing material may not be paper – water, maize cobs and leaves are common alternatives
BOD than richer people (Campos and von Sperling, 1996*)1 (further details
are given in Chapter 7). This is undoubtedly true in all developing countries,
but currently data only exist from Brazil.
Domestic wastewaters are collected in underground pipes which are called
‘sewers’. The flow in sewers is normally by gravity, with pumped mains only
being used when unavoidable.
The design of conventional sewerage (the sewer system used in
industrialized countries and in the central areas of many cities in developing
countries) is described in several texts (eg Metcalf and Eddy, Inc, 1986) and is
detailed in national sewerage codes (eg for India, Ministry of Urban
Development, 1993). However, it is extremely expensive. A much lower cost
alternative, which is suitable for use in both poor and rich areas alike, is
‘simplified’ sewerage, sometimes called ‘condominial’ sewerage. The design of
simplified sewerage is fully detailed by Mara et al (2001a*).
Untreated wastewater causes major damage to the environment and to human
health. Almost always, therefore, wastewater should be treated in order to:
reduce the transmission of excreta-related diseases (Chapter 2)
reduce water pollution and the consequent damage to aquatic biota
(Chapter 4).
6 Domestic Wastewater Treatment in Developing Countries
Only if there is a very large available dilution (>500) in the receiving
watercourse can consideration be given to discharging untreated wastewater
(see Table 4.2). For example, the city of Manaus (population in 2000: 1.4
million) in the Amazon region of Brazil discharges its wastewater untreated
via a river outfall into the Rio Negro, a tributary of the River Amazon, which
has a flow of ~30,000 m3 per second. The available dilution is >>500 and
therefore the pollution induced is negligible.
In developing countries only a small proportion of the wastewater
produced by sewered communities is treated. In Latin America, for example,
less than 15 per cent of the wastewaters collected in sewered cities and towns
is treated prior to discharge (Pan American Health Organization, 2000). Often
the reason for the lack of wastewater treatment is financial, but it is also due
to an ignorance of low-cost wastewater treatment processes and of the
economic benefits of treated wastewater reuse (Chapters 21 and 22); and also
because too many decision-makers appear happy to accept the status quo: the
continued discharge of untreated wastewater with its resultant damage to the
environment and human health. Currently the global burden of excreta-related
disease is extremely high (Chapter 2). Over half the world’s rivers, lakes and
coastal waters are seriously polluted by untreated domestic, industrial and
agricultural wastewaters (United Nations Environment Programme, 2002*;
Beach, 2001*), and they contain high numbers of faecal bacteria (Ceballos et
al, 2003*). Effective wastewater treatment needs to be recognized, therefore,
as an environmental and human health imperative.
Developing country governments and their regulatory agencies, as well as local
authorities (which may be city or town councils, or specific wastewater
treatment authorities, or more generally water and sewerage authorities), need
to understand that domestic and other wastewaters require treatment before
discharge or, preferably, re-use in agriculture and/or aquaculture. They also
need to act, but first they need to decide where, when and how much to invest
in wastewater treatment (Mariño and Boland, 1999*). Advice on the economic
analysis of investment projects is given by the World Bank (1996*; see also
Kalbermatten et al, 1982*).
Wastewater treatment for re-use in agriculture and aquaculture can be
subjected to classical benefit–cost analysis using discounted cash-flow
techniques to show if the present value of future additional crop yields is more
than the present value of wastewater treatment. However, wastewater
treatment prior to discharge to inland or coastal waters is less easy to analyse.
Central government, with its national perspective, must set national
environmental and environmental health priorities. It can enforce these by
lending money only for wastewater treatment projects that lie within these
priorities. Local authorities can then apply for a loan for a ‘priority’
wastewater treatment project. Generally, and ideally, priority projects should
What is Domestic Wastewater and Why Treat It? 7
be dealt with on the basis of river basin catchment areas, as this is the best
method of integrated water resources management, with central government
deciding which river basin is (or which river basins are) to be protected first,
what level of protection is needed now and how this can be developed to
progressively higher levels of protection in the future.
Wastewater treatment is needed on a truly enormous scale in developing
countries, and the purpose of this book is to show how it can be done at low
cost, and how treated wastewaters can be profitably and safely used in
agriculture and aquaculture – for wastewaters are simply too valuable to
An asterix after the year in a reference indicates that the publication referred to is
available on the Internet – see References.
Excreta-related Diseases
As noted in Chapter 1, one of the principal aims of domestic wastewater
treatment in developing countries is to reduce the numbers of excreted
pathogens to levels where the risks of further environmental transmission of
the diseases they cause are substantially reduced. Wastewater treatment
processes that are especially suitable for use in developing countries, such as
waste stabilization ponds (Chapters 9–13), are often designed specifically for
excreted pathogen removal. Wastewater treatment plant designers need,
therefore, to have a good understanding of excreta-related diseases, the
pathogens that cause them and how the plants they design can remove them.
A simple list of the 50 or so excreta-related diseases is not helpful to engineers,
nor is one which divides the list into viral, bacterial, protozoan and helminthic
diseases. What engineers (and other non-medical professionals) need is a list
that organizes the excreta-related diseases into categories according to their
environmental transmission route. This type of classification is called an
‘environmental’ classification, and this chapter presents the environmental
classification of excreta-related diseases developed in the early 1980s by
Professor Richard Feachem and his co-workers, mostly at the London School
of Hygiene and Tropical Medicine (Feachem et al, 1983*). In this chapter
Feachem’s classification has been annotated for use by wastewater treatment
and re-use engineers.
Table 2.1 gives an overview of Feachem’s environmental classification of
excreta-related diseases. There are seven categories (originally Feachem et al
had six; Mara and Alabaster, 1995, added the seventh). The first five comprise
the excreted infections – those in which pathogens in the excreta of one person
infect another person or persons. The last two categories are the vector-borne
excreta-related diseases – those excreta-related diseases spread by insects and
Excreta-related Diseases 9
Table 2.1 Environmental Classification of Excreta-related Diseases
I Non-bacterial
faeco-oral diseases
Low to medium
Unable to multiply
High infectivity
No intermediate host
Major examples
of infection
Hepatitis A and E
Rotavirus diarrhoea
Norovirus diarrhoea
II Bacterial
faeco-oral diseases Medium to high
Able to multiply
Escherichia coli
Medium to low
No intermediate host Shigellosis
III Geohelminthiases Latent
Very persistent
Hookworm infection
Unable to multiply
No intermediate host Trichurasis
Very high infectivity
IV Taeniases
Able to multiply
Very high infectivity
Cow or pig
intermediate host
V Water-based
Able to multiply
High infectivity
Intermediate aquatic
VI Excreta-related
insect-vector disease
filariasis transmitted
by Culex
VII Excreta-related
rodent-vector disease
transmission focus
Fodder crops
Aquatic species or
aquatic vegetables
Note: For medical details of all the diseases mentioned see, for example, Chin (2000) and Cook
and Zumla (2002)
10 Domestic Wastewater Treatment in Developing Countries
Excreted infections
The successful transmission of an excreted infection depends on the following
excreted load
The first five of these are properties of the pathogen, and the last is a property
of a potential host (ie the next person in the transmission chain).
Excreted load
Excreted load is the number of pathogens excreted by an infected person, and
it varies widely. For example, a person with cholera (a Category II disease)
may excrete ~1013 cholera vibrios per day. Someone with a light infection of
Ascaris lumbricoides, the human roundworm (Category III), may excrete a
few hundred thousand eggs per day (each female worm can produce up to
~200,000 eggs per day).
The excreted load depends on the state of infection: as the cholera victim,
for example, becomes better, the number of vibrios excreted falls – eventually,
of course, to zero. Another good example is schistosomiasis (Category V):
infected children generally show few clinical signs of the disease but excrete
large numbers of schistosome eggs, whereas adults in the terminal stage of the
disease excrete very few or no eggs.
The number of excreted pathogens in a wastewater depends on the number
of pathogens excreted by infected individuals in the community producing the
wastewater. Generally, numbers of endemic excreted pathogens in wastewater
are a few hundreds or thousands, occasionally tens of thousands, per unit
volume of wastewater considered (generally 100 ml for excreted bacteria, 1 l
for excreted worm eggs and protozoan cysts, and 10 l for excreted viruses –
see Chapters 11 and 12).
This is the interval between the excretion of a pathogen and it becoming
infective to another person or persons. Many excreted pathogens, including all
viruses, bacteria and protozoa (except Cyclospora), are non-latent: that is,
they are infective immediately upon excretion. Latency is an important
property, therefore, only of the helminths, and all the excreted helminths of
importance in wastewater treatment and re-use are latent. Their latency varies
from a few days to a few weeks, and during this time the worm changes from
a non-infective form to its infective form. This development may occur wholly
in the environment outside the body, as with the geohelminths (Category III),
Excreta-related Diseases 11
or it may take place partly in the environment and partly in an intermediate
host – a cow or pig in the case of the tapeworms (Category IV), or a water
snail and possibly also a fish or an aquatic vegetable in the case of the waterbased trematode worms (Category V).
How long an excreted pathogen can survive in the environment outside the
body is the property most indicative of the health hazard it poses. A pathogen
that is very persistent – for example, Ascaris eggs, which can survive for many
months, even years – are a risk in wastewater treatment and re-use. Even
excreted bacteria, which generally survive for only a few weeks, also constitute
a risk in this way.
Some excreted pathogens can, given the right environmental conditions,
multiply in the environment several thousand-fold or several million-fold: for
example, excreted bacteria in food and milk, and the water-based trematodes
in aquatic snails. Thus, a low excreted load can rapidly multiply to increase
the risk of infection. Excreted viruses and excreted protozoa cannot multiply
and, therefore, for them to be able to be transmitted successfully their
infectivity has to be very high.
Knowledge about infectivity – the probability of infection from one organism
– is far from perfect. What information there is has usually come from
volunteer studies: known pathogen doses are given to groups of volunteers
who are then monitored to see if they become infected (and, if they do, they
are then quickly treated). Generally, the volunteers have been healthy adults
from non-endemic areas, and their response is very different from that of
malnourished children in developing countries. Nevertheless, we cannot ignore
infectivity, however imperfect our knowledge. In general terms, we use the
following descriptive categories for the probability of infection from one
High infectivity
Medium infectivity
Low infectivity
Quantitative microbial risk assessment (Chapter 21) is used to calculate the
risks of infection and disease that may be associated with wastewater re-use.
Provided the wastewater has been well treated – specifically for pathogen
removal, especially in maturation ponds (Chapter 12) – these risks are very
low indeed.
12 Domestic Wastewater Treatment in Developing Countries
The excreta of one person will cause disease in an infected person, but only if
that person is susceptible. Host susceptibility governs the severity of the
disease: a person may be susceptible or, due to immunization or previous
exposure, be immune or have a varying degree of resistance.
Definition of terms
The terms ‘infection’ and ‘disease’ are often used interchangeably, but strictly
they have distinct meanings – an infected person may, or may not, become
diseased, depending on his or her susceptibility to the disease in question.
Incidence and prevalence are two important terms. Incidence is the number
of new cases of a particular disease in a community that occurs in a specified
time period (a week, a month or a year, but there is no relationship between,
say, weekly and annual incidences as all cases in a given year may have only
occurred in a given week). Incidence is generally used for acute diseases (eg
those in Categories I and II below). Prevalence is the number (or proportion)
of people in a community with a particular disease at a specified point in time.
It is used for chronic infections or diseases such as the various excreta-related
helminthiases (Categories III–V below). It is possible to refer to the incidence
of one of these chronic diseases, but this has the meaning given above (and is
useful, for example, when determining reinfection following community-wide
antihelminthic chemotherapy).
Category I: Non-bacterial faeco-oral diseases
The term faeco-oral is used to describe the beginning and end of the excreted
pathogen’s transmission route: it leaves one person in his or her faeces and
enters another person through his or her mouth. This category includes all the
excreted viral and protozoan diseases, and these excreted pathogens are nonlatent, have a low-to-medium persistence, are unable to multiply, have a high
infectivity and do not have an intermediate host. These infections are mainly
spread in a very direct person-to-person way wherever personal and domestic
hygiene is poor. However, those that can survive for several days (the protozoa,
for example) are also important in wastewater treatment and re-use.
The most important viruses in this category are rotaviruses and
noroviruses (until recently the latter were called Norwalk and Norwalk-like
viruses), which are the principal causes of viral diarrhoea in both developing
and industrialized countries. Rotaviruses cause 350,000–600,000 deaths per
year in children under five years old, 82 per cent of which are in developing
countries (Parashar et al, 2003*). Other important diarrhoeagenic viruses are
adenoviruses, astroviruses and other caliciviruses.
There are four main protozoa that cause diarrhoea: Entamoeba histolytica,
Giardia intestinalis (also called G lamblia), Cryptosporidium parvum and
Cyclospora cayentanensis. The first three are non-latent, whereas Cyclospora
is latent and requires a period of seven to ten days to sporulate into its
infectious form (Relman, 1998). Outbreaks of cyclosporiasis in the US and
Canada during 1996–2000, associated with the consumption of Guatemalan
Excreta-related Diseases 13
raspberries (see Ho et al, 2002*), resulted in a ban on their import into the US
and consequent huge economic losses and unemployment in Guatemala. The
prevalence of cyclosporiasis amongst Guatemalan raspberry farm workers,
especially children under the age of ten, was higher than in non-farm workers
(Bern et al, 2000*). There was an outbreak of cyclosporiasis in south-west
Germany in 2000–2001 associated with the consumption in a restaurant of
salad side dishes prepared with lettuce imported from southern France and
herbs and spring onions from southern Italy (Döller et al, 2002*).
Category II: Bacterial faeco-oral diseases
The transmission features of the bacterial excreted pathogens are that they are
non-latent, have a medium-to-high persistence, are able to multiply, have a
medium-to-low infectivity and do not have an intermediate host. These
infections can be transmitted in the same direct person-to-person way as
Category I infections, but their greater persistence means that they are even
more important in wastewater treatment and re-use.
The major excreted bacterial pathogens are Campylobacter spp,
diarrhoeagenic E coli, Salmonella spp, Shigella spp and Vibrio cholerae. Most
of the global incidence of bacterial diarrhoea is associated with Campylobacter
and diarrhoeagenic E coli. The two species of Campylobacter pathogenic to
humans are Campylobacter jejuni and Campylobacter coli, and they are often
present in waters and wastewaters (Jones, 2001*). Children are most at risk,
especially those under the age of two years, in whom polymicrobial infection
is common (ie infection with both Campylobacter and one or more other
gastrointestinal pathogens) (Coker et al, 2002*). Guillain–Barré syndrome (the
most common form of acute neuromuscular paralysis) is a potential severe
outcome of C jejuni infection (which is the most usual cause, although it can
be induced, but generally at lower severity, by other non-excreted bacteria and
viruses) (Hadden and Gregson, 2001*).
Most E coli strains are non-pathogenic commensal inhabitants of the
gastro-intestinal tract of humans and most animals. However, diarrhoeagenic
E coli strains are extremely pathogenic; they comprise several types termed
(mainly after their pathogenesis) enterotoxigenic E coli (or ETEC),
enteropathogenic E coli (EPEC), enterohaemorrhagic E coli (EHEC),
enteroaggregative E coli (EAEC), enteroinvasive E coli (EIEC) and diffusively
adhesive E coli (DAEC) (Nataro and Kaper, 1998*; Hunter, 2003*). ETEC is
a very common pathogen (see Chart, 1998*) and is the second most important
bacterial cause of diarrhoea after Campylobacter. EHEC includes E coli O157,
a virulent serotype causing high mortality in the most vulnerable groups (the
very old and the very young).
Categories I and II are very similar, the only difference being the greater
persistence of the excreted bacteria. Categories III–IV are very different: they
comprise the excreted helminthic infections, the pathogens are all latent and
very persistent, and some have one or more intermediate hosts. An excellent
introduction to helminthic diseases is given by Muller (2001*).
14 Domestic Wastewater Treatment in Developing Countries
Category III: Geohelminthiases
This category contains the geohelminths – the soil-transmitted nematode
worms. The main ones of importance in wastewater treatment and re-use are:
Ascaris lumbricoides – the human roundworm
Trichuris trichiura – the human whipworm
Ancylostoma duodenale and Necator americanus – the human
Their transmission features are that they are latent, very persistent, unable to
multiply, have a very high infectivity and do not have an intermediate host.
They are extremely common pathogens, especially Ascaris and the
hookworms. In low-income areas of developing countries prevalences are often
over 50 per cent (ie over half the population is infected), and prevalences
greater than 90 per cent occur frequently. The number of worms per person
(the ‘worm burden’, a measure of the intensity of infection) can also be high
(Figure 2.1).
Adult female Ascaris worms each produce up to ~200,000 eggs/day, which
leave the body in the faeces, and adult female hookworms each produce
5000–20,000 eggs/day, which also leave in the faeces. Egg numbers in
wastewater can thus be quite high in endemic areas, up to ~3000/l. Fortunately
they are easily removed in several of the wastewater treatment processes
described in this book (Chapters 9–19), and thus compliance with the World
Health Organization’s guideline value of ≤1 egg/l treated wastewater used for
crop irrigation, or ≤0.1 egg/l when children under 15 are exposed usually is
not a problem (World Health Organization, 1989*, 2004*; see also
Blumenthal et al, 2000*); further details are given in Chapters 12 and 21.
Category IV: Taeniases
This category contains the two main human cestode worms: Taenia saginata,
the beef tapeworm, and Taenia solium, the pork tapeworm. Their transmission
features are that they are latent, persistent, able to multiply, have a very high
infectivity and have a cow or pig intermediate host. Around 105–106 eggs are
produced per day by each worm, and these leave the body in the faeces inside
gravid segments of the worm, from which the eggs are released into
wastewater. Taenia eggs are also easily removed in wastewater treatment
The embryonic form of T solium (the ‘cysticercus’) can enter the brain
where it may induce neurocysticercosis. This is the leading cause of epilepsy in
developing countries (except Muslim and other countries or communities
where pork is not eaten) (Sotelo, 2003*).
Category V: Water-based helminthiases
This category contains all the water-based human trematode worms. There are
several of these, but only three genera are of major importance:
Excreta-related Diseases 15
Figure 2.1 This 4-year old African Girl with a Distended Abdomen was
Given an Appropriate Dose of the Vermifuge Levamisol; Shortly Afterwards
she Excreted the Large Number of Adult Ascaris lumbricoides Worms Shown
16 Domestic Wastewater Treatment in Developing Countries
Schistosoma mansoni, S japonicum and S haematobium, the main human
schistosomes or blood flukes
Clonorchis sinensis, the oriental liver fluke (found mainly in China, Japan,
Korea and Vietnam)
Fasciolopsis buski, the giant human intestinal fluke (found mainly in India,
Bangladesh, Thailand, Cambodia, China, Malaysia, Indonesia, Vietnam,
Laos and the Philippines).
Their transmission features are that they are latent, reasonably persistent, able
to multiply, have a high infectivity, and have one or two intermediate aquatic
hosts – a snail (all three) and then either fish (C sinensis) or aquatic vegetables
(F buski).
Adult female schistosomes produce up to ~1000 eggs/day, adult female
Clonorchis worms up to ~4000 eggs/day, and adult female Fasciolopsis worms
~25,000 eggs/day. The eggs are voided in faeces (or, in the case of
S haematobium, in urine), and they hatch in wastewater to form miracidia,
which then have to enter a specific species of water snail in order to continue
their life cycle. These trematode infections are potentially important in the
aquacultural re-use of treated wastewaters (Chapter 22), but they are easily
removed during wastewater treatment.
Category VI: Excreta-related insect-vector diseases
The only disease in this category important in wastewater treatment and reuse is Bancroftian filariasis when it is transmitted by the mosquito Culex
quinquefasciatus, which can breed in poorly maintained wastewater treatment
plants (Chapter 14). It is a serious disease caused by the nematode worm
Wuchereria bancrofti. Adult worms live in the lymphatic ducts of humans, and
embryo worms (called ‘microfilariae’) are shed in large numbers into the
bloodstream at night. If a culicine mosquito ingests microfilariae during its
blood meal, they develop inside the mosquito over a period of 10–15 days to
become infective larvae. When the mosquito feeds again, they are introduced
into another person where they develop over 3–12 months into adult male and
female worms that establish themselves in the lymphatic system, and the cycle
of microfilarial production recommences. After a few years of infection the
lymph glands and lymphatic vessels become partially blocked and swollen as
the lymph cannot drain. This leads to swelling of the genitalia, legs or arms,
and the resulting gross deformity is called ‘elephantiasis’. Bancroftian filariasis
is becoming increasingly common in urban areas that have good water supplies
but poor sanitation – the resulting wastewaters pond in garbage-blocked
stormwater drains and natural drainage channels, so permitting the culicine
vector mosquitoes to proliferate. The solution is to install low-cost simplified
sewerage (Mara et al, 2001a*) and a properly designed wastewater treatment
plant that is well-operated and maintained.
Excreta-related Diseases 17
Category VII: Excreta-related rodent-vector diseases
The only disease in this category of relevance in wastewater treatment and reuse is leptospirosis, which is caused by the bacterium Leptospira interrogans.
Leptospirosis is primarily a disease of brown rats, and humans become infected
when they come into contact with infected rat urine. The leptospires then enter
the body through damaged skin (a cut or abrasion). In humans the infection
can be asymptomatic with mild (influenza-like) symptoms, or severe – the
severest form is Weil’s disease, and this can be rapidly fatal if not treated;
symptoms include jaundice – skin and eye haemorrhages, and liver and kidney
failure. Sewer maintenance workers are especially at risk, but the disease is
also a potential risk in wastewater treatment and re-use. It is a becoming a
more common infection in India, for example (Chaudhry et al, 2002*).
Emerging infectious diseases
Many excreta-related diseases are ‘new’ diseases – in the sense that their
causative agent is a newly discovered pathogen. These diseases are termed
‘emerging’ infections, and several very important excreta-related pathogens
have only been discovered in the last 30 years (Table 2.2). Some diseases are
‘re-emerging’ as, due to changing circumstances (eg a high HIV/AIDS
prevalence), the pathogens are now able to infect more people more frequently.
Unquestionably, more emerging excreta-related pathogens will be found, but it
should be relatively easy to assign them to the appropriate category in the
environmental classification of excreta-related diseases given above.
Excreta-related cancers
Long-term infection with some excreted pathogens can induce cancer – for
example, the water-related helminths (Category V): bladder cancer is induced
by Schistosoma haematobium, colorectal cancer by S mansoni and
S japonicum, and bile duct cancer by Clonorchis sinensis (Mara and Clapham,
1997*). Helicobacter pylori, a faeco-oral (and also oro-oral) bacterial
pathogen that causes stomach ulcers, can induce stomach cancer. It is
Table 2.2 Major Excreta-related Pathogens Identified Since 1973
Hepatitis A virus
Cryptosporidium parvum
Campylobacter spp
Cyclospora cayetanensis
Escherichia coli O157
Helicobacter pylori
Hepatitis E virus
Vibrio cholerae O139
Source: Satcher (1995*); Favorov and Margolis (1999)
18 Domestic Wastewater Treatment in Developing Countries
extremely common, with infection prevalences of 50–80 per cent in developing
countries, and it is the only bacterium to be designated as a known human
carcinogen (International Agency for Research on Cancer, 1994; see also
Engstrand, 2001*; Frenck and Clemens, 2003*).
A ‘snapshot’ of the global burden of all diseases in 1990 is given by the World
Bank–World Health Organization study conducted by the Harvard School of
Public Health (Murray and Lopez, 1996a, 1996b). Table 2.3 lists the global
incidence of diarrhoeal diseases and the global prevalence of geohelminthic
infections in 1990, essentially all of those which occurred in developing
countries. Indeed, diseases due to deficient water supplies, deficient sanitation
and deficient hygiene were together responsible for 7 per cent of all deaths in
the world in 1990, second only to malnutrition, which caused 15 per cent of
all deaths (Murray and Lopez, 1996a). However, by 2000 the proportion of
deaths due to these diseases had fallen to 4 per cent (Prüss et al, 2002*).
(Deaths from HIV/AIDS are increasing rapidly, especially in Africa – see The
Lancet, 2002*, and are likely to overtake deaths due to deficient water
sanitation and hygiene soon.)
The incidence of excreta-related diseases shows little sign of decline,
especially in developing countries, but also in industrialized countries – for
example, in England the annual incidence of infectious intestinal disease from
all causes (but mainly food-borne diarrhoea) is 0.2 per person (Wheeler et al,
1999*), much lower than the incidence of diarrhoea in developing countries
(1.3/person/year, Table 2.3), but even so is very high. In developing countries
diarrhoea is still a major killer: some 1.3 million children under the age of five
Table 2.3 Global Diarrhoeal Disease and Geohelminthiases Statistics for
4,073,920,000 episodes
56% in children aged 0–4
94% in developing countries
61,847,000 persons with
high-intensity infection
73% in children aged 5–14
All in developing countries
45,421,000 persons with
high-intensity infection
79% in children aged 5–14
All in developing countries
152,492,000 persons with
high-intensity infection
36,014,000 persons with
84% in adults aged 15–59
All in developing countries
72% in adults aged 15–44
All in developing countries
Note: The world population in 1990 was 5.3 billions, of which 3.9 billions (74%) were in developing
Source: Murray and Lopez (1996b)
Excreta-related Diseases 19
die from it each year (ie one diarrhoeal-disease child death every 25 seconds).
More insidiously, diarrhoea in infancy is associated with ‘stunting’ (a medical
term for low height-for-age, ie impaired growth) and also with poor cognitive
function (ie impaired mental development) in later childhood (Berkman et al,
2002*). Children under the age of five years form only 10 per cent of the
world’s population, yet they bear at least 40 per cent of the total global burden
of environmental – including excreta-related – disease (World Health
Organization, 2002*).
Geohelminthic infections are extremely common: approximately one-third
of the world population – some 2 billion people – has intestinal worms (Chan,
1997*). These worms eat their food before they do, so contributing to
malnutrition and hence retarded growth and impaired cognition. In the case of
the human hookworms, which hook into their hosts’ stomach wall and drink
their blood, anaemia is common, and women of child-bearing age can lose
more blood in this way than through menstruation.
Wastewater treatment engineers, and more generally tropical public health
engineers, have an extremely important role in reducing the environmental
transmission of excreta-related diseases and, by so doing, in greatly
contributing to socio-economic development in developing countries. This
contribution is potentially enormous: Pearce and Warford (1993) quote data
for 1979 (with the implication that more recent data did not exist): in that
year some 360–400 billion working days in developing countries were lost
from water- and excreta-related diseases that kept people from work. Valuing
a working day lost at only US$0.50, these countries therefore lost US$180–200
billion in that year and, as the GNP of all developing countries was then
US$370 billion, output was below potential production by as much as 33–35
per cent. Of course, good wastewater treatment is only one of the means
tropical public health engineers have to combat excreta-related diseases, but it
is an important one. Treatment combined with productive re-use (Chapters 21
and 22) contributes even more directly to socio-economic development.
Essential Microbiology and Biology
Wastewater treatment and re-use engineers need a good understanding of
wastewater microbiology for two reasons: first because wastewaters contain
micro-organisms that cause human disease (Chapter 2), and secondly because
most wastewater treatment processes are microbiological (we generally use the
term ‘biological wastewater treatment’ to reflect this, but we should say
‘microbiological wastewater treatment’). Wastewater treatment engineers also
need to understand the effect of untreated, partially treated and fully treated
wastewaters on the biology – really, the aquatic ecology – of the receiving
This chapter provides an introduction to the important groups of microorganisms in wastewater treatment and reuse: viruses, bacteria, algae,
protozoa and helminths (more detailed information is given in, for example,
Mara and Horan, 2003). It concludes with a brief description of a simplified
technique for the biological assessment of tropical freshwater quality using
aquatic micro-invertebrates.
What are micro-organisms?
Micro-organisms (often simply called ‘microbes’) are small single-celled
organisms: viruses, bacteria, micro-algae and protozoa. To see them we need
to use a microscope. (Helminths are multicellular animals and therefore not
microbes, but we include them in the general topic of tropical sanitary
microbiology, although this should be more properly described as tropical
sanitary microbiology and parasitology.)
Micro-organisms are the ancestors of all organisms that exist (or have
existed) on Earth, and they are the most numerous of all organisms: a handful
of soil contains many hundreds of billions of microbes, mostly bacteria – more
than the world’s stock of plants and animals (‘macro-organisms’). Microorganisms are relatively simple life forms, but it is a mistake to think of them
as ‘primitive’ – they are complex biochemical ‘machines’ which serve us well
(they are the prime movers in the biogeochemical cycles of oxygen, carbon and
nitrogen, for example, without which no life would exist), but which can also
Essential Microbiology and Biology 21
serve us badly (the Spanish influenza pandemic of 1918–1919, for example,
killed ~40 million people; Brainerd and Siegler, 2003*). Fortunately only a few
micro-organisms serve us badly.
The naming of organisms follows strict international rules. Each organism is
given two Latin names (both of which are written in italics): the first (which
always commences with a capital letter) denotes the organism’s genus, and the
second (which does not) its species. Thus the common mammalian gut
bacterium Escherichia coli belongs to the genus Escherichia and its specific
epithet (ie species name) is coli. The organism is often referred to as E coli (ie
the generic name is abbreviated to its initial letter, sometimes to its initial two
or occasionally three letters to avoid confusion). Similar genera are grouped
into families. Escherichia coli belongs to a family of gut bacteria called the
Enterobacteriaceae (bacterial family names have the distinctive ending ‘aceae’,
and are not italicized).
We can refer only to the genus – for example, Escherichia. We can also
refer to an unnamed species of this genus as Escherichia sp, or to more than
one unnamed species as Escherichia spp. More informally, we can refer to
some important (usually medically important) genera as, for example,
salmonellae (for Salmonella spp), shigellae (Shigella spp) and vibrios (Vibrio
Singulars, plurals and adjectival forms
The singular, plural and adjectival form of types of microbes are mainly Latin
or Greek, as follows:
Bacterium, bacteria; bacterial
Virus, viruses; viral
Protozoon, protozoa; protozoan
Alga, algae; algal.
It may seem pedantic to introduce these points of grammar, but few engineers
use these terms correctly (and there is no real excuse for this – not even
‘microbiological ignorance’).
Domains of life
Modern evolutionary biologists now consider a ‘Tree of Life’ with three
‘Domains’: the Bacteria, the Archaea and the Eukarya (Woese et al, 1990; see
also Gupta, 2000*). This three-domain paradigm (Figure 3.1) shows how all
living organisms derived from a single common ancestor (long since extinct or
at least as yet undiscovered). Both the Archaea and the Bacteria are
prokaryotes – that is, they are single-celled organisms which do not have a
clearly defined cell nucleus. The first cell (or cell-like entity) emerged more
than 4 billion years ago, and the Bacteria developed from the ‘last universal
22 Domestic Wastewater Treatment in Developing Countries
ancestor’ (ie a descendent of the first cell) roughly 3 billion years ago. The
evolutionary development that set the Bacteria apart from these earlier cells
was that of an essentially rigid, completely enclosing, external cell wall (termed
a ‘sacculus’, basically the microbial equivalent of an exoskeleton) which
enabled them to withstand a high internal osmotic pressure (ie it prevented the
cells from rupturing) (Koch, 2003). The Archaea evolved from the last
common ancestor (or one of its non-Bacterial descendents) later; they are a
very important group for wastewater engineers as they include the methane
producers. The Archaea and the Bacteria thus comprise the micro-organisms
that we have for the past 120 years or so called ‘bacteria’.
All other life forms (ie all fungi, plants and animals – including ourselves)
are eukaryotes (ie their cells have a clearly defined nucleus) and they form the
third domain: the Eukarya, which also evolved from the last common ancestor,
probably after the Archaea. The domain Eukarya includes the familiar
kingdoms of Animalia, Plantae and Fungi (with the helminths within the
Animalia), and also the green algae and the protozoa.
This modern view of the Tree of Life emphasizes the huge importance of
the microbial world, not only in sheer numbers but also in the ecology of Earth
and in the evolution of Life.
The microbes of importance to wastewater treatment engineers come
mainly from the Bacteria, some from the Archaea, and some from the Eukarya
(eg the pathogenic protozoa – Chapter 2; and the green algae that are the
‘workhorses’ of facultative and maturation waste stabilization ponds –
Chapters 9, 11 and 12). Some higher Eukarya are also important – for example
the pathogenic helminths (Chapters 2 and 21) and the plants used in
constructed wetlands (Chapter 17).
Good introductory texts on microbiology are Heritage et al (1996, 1999).
Suitable reference texts on environmental microbiology are Hurst et al (2001)
and Bitton (2002), and on water and wastewater microbiology Water
Environment Federation (2001) and Mara and Horan (2003).
Viruses are extremely small (~20–200 nm) parasitic microbes which can
reproduce only by invading a host cell whose reproductive processes they
redirect to manufacture more viruses. The structure of viruses is extremely
simple: they comprise a core of either DNA or RNA surrounded by a protein
‘coat’ (or ‘envelope’).
Once an infectious virus enters a host cell, the virus replicates itself
hundreds or thousands of times; these new viruses leave the host cell and in so
doing destroy it – and it is the death of these cells that causes disease in the
host. Viruses are mostly very specific in their choice of host – plant viruses
cannot invade animal cells, for example. Domestic wastewater contains many
human viruses, including the rotaviruses and noroviruses that are the major
viral causes of diarrhoea (Chapter 2).
Essential Microbiology and Biology 23
~3 billion years ago
~4 billion years ago
Note: This is a very simplified version which highlights the organisms of importance in wastewater
treatment. The branch locations in the bacterial and archaeal domains are not chronological or
even indicative as the relative time at which the different groups evolved is not known with any
certainty. For simplicity the bacterial domain, in particular, is presented in only the barest outline.
The Gram-negative phylum Proteobacteria has five classes (alpha to epsilon) and contains a very
large number of the micro-organisms of importance to wastewater treatment engineers – for
example, the enteric bacteria (coliforms, E coli, salmonellae, shigellae, vibrios, etc), most of the
important chemoheterotrophs (eg the ‘BOD (biochemical oxygen demand) removers’) and
chemo-autotrophs (including most of the nitrifiers), and some phototrophs (the purple bacteria,
the Chromatiaceae). Other important bacterial phototrophs are the Cyanobacteria and the
Chlorobi (containing the Chlorobiaceae). The phylum Bacteroidetes, also Gram-negative,
contains the Bacteroides-like anaerobes and the Cytophaga–Flavobacterium complex of aerobic
heterotrophs. There are two main phyla of Gram-positive bacteria: the Firmicutes (including
Bacillus spp and Clostridium spp) and the Actinobacteria (eg Nocardia spp, implicated in
Figure 3.1 The Tree of Life with its three Domains
Bacterial viruses are called ‘bacteriophages’ and these can be used to model
viral die-off in waste stabilization ponds (Chapter 12).
Viruses do not fit into the three-domain Tree of Life shown in Figure 3.1 –
their evolutionary position is not yet fully understood. They may possibly be
prokaryotes that have evolved from intracellular parasitic forms (ie
prokaryotes that could only reproduce inside the cells of a eukaryotic or
prokaryotic host) to become the ‘simple’ packets of RNA or DNA that they
are today. (There is one other group whose evolutionary position is also not
understood: the prions, which are protinaceous infectious particles that cause
diseases such as the spongiform encephalopathies – scrapie in sheep, ‘mad cow
24 Domestic Wastewater Treatment in Developing Countries
disease’ and human Creutzfeldt-Jacob diseases, including ‘kuru’ and new
variant CJD.)
Although the Bacteria and Archaea are as different, in evolutionary terms,
from each other as they both are from the Eukarya (Figure 3.1), they are
considered together in this section since they are both prokaryotic and their
growth follows the same rules. In general (and unless otherwise indicated) the
terms ‘bacterium’, ‘bacteria’ and ‘bacterial’ are used in this book to refer to
both the Bacteria and the Archaea – strictly speaking, of course, it would be
better to refer to them as prokaryotes (but sometimes scientific exactitude can
be too confusing, or at least too obfuscatory).
Shape, size and structure
Bacteria and Archaea are small, just a few micrometres (µm) in size, and their
mass is around a picogram (pg – ie 10–12 g), although some ‘giant’ bacteria do
exist (for example, Thiomargarita namibiensis is 100–300 µm in diameter;
Schulz, 2002), and they come in several different shapes (Figure 3.2). To see
them we need to use a microscope; for our engineering purposes a light
microscope is sufficient, with an oil-immersion x100 objective lens and a x10
eye-piece to give a magnification of x1000. Unless we use a phase-contrast
microscope, we need to stain the bacteria before looking at them under the
microscope. The staining procedure most commonly used is the Gram stain
(devised in 1884 by Hans Christian Gram, a Danish bacteriologist); illustrated
details of the procedure are given by the University of Leicester (2002*). Grampositive bacteria appear purple and Gram-negative bacteria red when stained
according to this procedure. This difference is very important, not just as a
staining technique, but because Gram-positive and Gram-negative bacteria are
different in a much more fundamental way: Gram-positive bacteria have a cell
wall structure comprising a single membrane (so they are called ‘monoderms’),
whereas Gram-negative bacteria have a double-membrane cell wall structure
(‘diderms’). All Archaea are Gram-positive and so, in evolutionary terms,
Gram-positive Bacteria are closer to the Archaea than Gram-negative Bacteria
are (Gupta, 2000*).
Environmental requirements
Bacteria vary widely in their environmental requirements and preferences. For
example, some bacteria can only grow in the presence of oxygen – the ‘obligate
aerobes’; some can only grow in its absence – the ‘obligate anaerobes’; and
some can grow in both its presence and in its absence, although growth is
better in its presence – the ‘facultative anaerobes’ (or simply ‘facultative
bacteria’). Most bacteria cannot use carbon dioxide as a source of cell carbon,
but some can (and prefer to do so); those that cannot are termed ‘heterotrophs’
Essential Microbiology and Biology 25
Figure 3.2 Common Bacterial Shapes (cocci are ~1 µm in diameter, and
bacilli are typically 1 x 3–6 µm)
and those that can are ‘autotrophs’. Some bacteria can photosynthesize, and
these are either ‘photoheterotrophs’ or ‘photo-autotrophs’; however, most
bacteria cannot photosynthesize and they are therefore either
‘chemoheterotrophs’ or ‘chemo-autotrophs’.
Temperature is a very important environmental parameter. Most bacteria
grow well in the temperature range 15–40 degC and are termed ‘mesophils’;
some grow best at lower temperatures – the ‘psychrophils’; and some require
much higher temperatures (some even close to the boiling point of water) – the
‘thermophils’. In wastewaters in tropical and subtropical regions most bacteria
are, as would be expected, mesophilic.
The pH of the environment in which bacteria grow is another important
environmental parameter. Most bacteria prefer near neutral or slightly alkaline
conditions, around pH 6.5–8.5; some can tolerate pH >9 (eg Vibrio cholerae,
the causative agent of cholera); and some generate very acid conditions (eg
Thiobacillus thioparus which produces sulphuric acid at pH <2 and so causes
rapid sewer crown corrosion in warm climates).
Salt is an environmental parameter of importance, for wastewater
treatment engineers, only in that freshwater and faecal bacteria cannot grow
in very saline waters (the sea, for example; this is relevant when treated
wastewaters are discharged into coastal waters – Chapter 4). Marine bacteria,
in contrast, are ‘halophils’.
Domestic wastewater fortunately contains roughly the right balance of
nutrients for bacterial growth – a BOD:N:P ratio of ~100:5:1. The presence of
industrial effluents can alter this ratio and the wastewater may need nitrogen
and/or phosphorus supplements.
26 Domestic Wastewater Treatment in Developing Countries
Bacterial growth kinetics
Bacteria grow by binary fission: a cell divides into two daughter cells. These
grow and each divides into two more cells; so the sequence of cell numbers
originating from one cell is 1, 2, 4, 8, 16 and so on, reaching 2n after n
divisions. The rate of growth (ie the number of divisions per unit time) depends
on many factors in the immediate environment of the dividing cells, as
described above. Often one factor is growth-limiting; this could be oxygen (for
aerobes), temperature (which is why we store food in a refrigerator), a suitable
source of carbon (or nitrogen or phosphorus, or an essential micronutrient
such as a vitamin), or too low or too high a pH.
If a bacterium takes T minutes to divide into two, then it takes nT minutes
to multiply to 2n cells. This type of growth is logarithmic (or exponential) and
it is described by the equation:
Nt = N0exp(µt)
where Nt and N0 are the number of cells present at time t (minutes) and
initially, respectively; and µ is the specific growth rate, which has units of
reciprocal time (usually day–1, but here minute–1). The term ‘exp’ means ‘e to
the power of what follows in brackets, where e is the base of Naperian
logarithms (denoted ‘ln’) and equals ~2.7183.
If t = T (which is the doubling, or mean generation, time), then from
equation 3.1 with Nt = 2N0:
µ = (ln2)/T = 0.69/T
The bacterial growth curve
If we introduce a few bacteria into, say, a litre of soluble waste, and if no
further additions of waste are made, the bacteria will typically exhibit four
distinct phases of growth (Figure 3.3). The first is the lag phase, during which
time cell numbers do not increase; the bacteria are, however, internally active,
manufacturing if necessary any intracellular catalysts (‘enzymes’) that they
may require in order to be able to oxidize the waste. Next comes the
exponential phase in which logarithmic growth occurs; during this phase the
bacteria lay down food reserves within their cells which they may use when
there is little or no food left in their environment. The bacteria are now
growing as fast as they are able to in the waste; equation 3.1 is therefore
rewritten as:
N = N0exp(µmaxt)
where µmax is the maximum specific growth rate.
Essential Microbiology and Biology 27
3. stationary
log cell numbers
1. lag
Time, hours
Figure 3.3 The Bacterial Batch-culture Growth Curve (axis numbers are
illustrative only)
The exponential phase ceases, often abruptly, either because the supply of an
essential nutrient has been exhausted or because there has been an
accumulation of toxic end-products of bio-oxidation (an example of the latter
is the accumulation of acid that is an end-product of the bio-oxidation of
sugars; the pH falls to a growth-inhibiting level). In the ensuing stationary
phase the number of new cells is approximately balanced by those that die, so
that the cell population does not change. When the death rate exceeds the
growth rate, the culture enters the death phase and the population steadily
declines. During both the stationary and death phases there is a substantial
proportion of cells which neither die nor subdivide; they exist by utilizing the
intracellular food reserves laid down during exponential growth, a process
known as ‘endogenous respiration’. When a cell has depleted its food reserves,
it starts to oxidize itself; this process, known as ‘autolysis’ (ie self-destruction),
leads of course to death.
Continuous culture
The microbiological processes used for wastewater treatment operate
continuously, 24 hours a day for 7 days a week and 52 weeks a year, rather
than as a batch process. Bacterial growth in a continuous reactor occurs at a
rate less than the maximum growth in batch culture – that is µ <µmax. The
value of µ depends on the value of the growth-limiting substrate; this may
28 Domestic Wastewater Treatment in Developing Countries
often be the wastewater strength, expressed in terms of BOD (Chapter 1), but
it could be another parameter (such as ammonia, if nitrification is being
considered – see later in this chapter). The Monod equation is used to
determine the value of µ:
( L +LK ) – b
where L is the BOD in the reactor, mg/l; KL is the Monod ‘half saturation’
constant (= the value of L when µ = µmax/2), mg/l; and b is the endogenous
decay rate, which has the same units as µ, usually day–1.
First-order BOD removal kinetics (Chapter 5) are generally used for
microbiological reactor design, rather than equation 3.4, although the latter is
used in activated sludge design and also for nitrification in aerated lagoons
(Chapter 20). Monod kinetics have recently been applied to facultative waste
stabilization ponds (Kayombo et al, 2003*).
Equations 3.2 and 3.4 combined are important in reactor design, as they
give the minimum cell retention time in a continuous-flow microbiological
reactor at which cell growth is balanced by the number of cells leaving the
reactor. If the cell retention time is less than this, then the cells will be
exponentially ‘washed out’ of the reactor and reactor failure quickly ensues.
Wastewater treatment engineers have to ensure that the cell retention time in
the treatment units they design is longer than this minimum value.
Anabolism and catabolism
Bacteria oxidize wastes to provide themselves with sufficient energy to enable
them to synthesize the complex molecules such as proteins and polysaccharides
which are needed to build new cells. Thus bacterial metabolism has two
component parts: catabolism (‘breaking down’) for energy and anabolism
(‘building up’) for synthesis. The verbal ‘equation’
wastes + oxygen oxidized waste + new bacteria
is instructive but oversimplified in that the anabolic and catabolic reactions
are not distinguished; nor is there mention of autolysis, which is an important
form of catabolism. The following three equations describe these processes
CxHyOzN + O2 CO2 + H2O + NH3 + energy
Essential Microbiology and Biology 29
CxHyOzN + energy C5H7NO2 (ie bacterial cells)
C5H7NO2 + 5O2 5CO2 + NH3 + 2H2O + energy
As a general guide ~1/3 of the available BOD is used in catabolic reactions and
~2/3 in anabolic reactions (Figure 3.4). The equation for autolysis does not
proceed to completion since approximately 20–25 per cent of the cell mass is
resistant to bacterial degradation.
Bacteria in wastewater treatment
Wastewater contains many billions of bacteria, and most of these are faecal
bacteria. However, these bacteria, once outside their normal habitat (ie when
they are in the ‘extra-intestinal’ environment), are unable to survive for very
long. This is because they are outcompeted by the large numbers of
saprophytic bacteria which grow naturally and profusely in the nutrient-rich
aquatic environment of a microbiological wastewater treatment reactor. These
‘saprophytes’ obtain their energy, cell carbon and other essential nutrients from
the organic and inorganic compounds in the wastewater. They are very well
adapted to this environment, whereas the faecal bacteria are not – and so they
CO2 +H2O +energy
Units of BOD5
CO2 +H2O +energy
inert cell residue
(unmetabolized fraction)
Note: In a real (finite time) continuous microbiological reactor some of the organic matter (ie
BOD) in the influent escapes oxidation; in batch culture at infinite time the unmetabolized fraction
is zero.
Figure 3.4 The Catabolic, Anabolic and Autolytic Reactions of Aerobic
Microbiological Oxidation
30 Domestic Wastewater Treatment in Developing Countries
die, some quickly and some more slowly (the kinetics of faecal bacterial dieoff is important in the design of maturation ponds – Chapter 12).
The most commonly isolated saprophytic bacteria in aerobic
microbiological treatment systems are Gram-negative, facultative,
heterotrophic rods. They mostly belong to the genera Achromobacter, Bacillus,
Flavobacterium, Pseudomonas and Zooglea, together with the non-faecal
coliforms (see below); all these bacteria are Proteobacteria (Figure 3.1).
Nitrification is the oxidation of ammonium to nitrate, which is done by two
groups of obligately aerobic autotrophic proteobacteria. First ammonium is
oxidized to nitrite by the ammonium oxidizers – for example, Nitrosomonas
spp and Nitrospira spp:
NH4 + 1.5O2 NO2 + H2O + 2H+
The nitrite so produced is then oxidized to nitrate by the nitrate oxidizers,
typically Nitrobacter spp, with the oxygen atom added to the nitrite ion
coming from water (rather than from molecular oxygen):
NO2 + H2O NO3 (= NOO2) + 2H+
A more descriptive equation for overall nitrification, which shows the
formation of nitrifying bacterial cells (C5H7NO2), is the following (Horan,
NH+ + 1.83O2 + 1.98HCO– 0.021C5H7NO2 + 0.98NO– + 1.04H2O +
This shows that 1 mole of ammonium-N (ie 14 g N) requires 1.83 moles of
oxygen (58.6 g O2) for nitrification – that is the nitrification oxygen demand
is 4.2 g O2 per g ammonium-N. However, this equation does not take into
account the fact that the oxygen used in the oxidation of nitrite to nitrate
comes from water, not molecular oxygen. To allow for this, 0.98H2O must be
added to each side of the equation:
NH+ + 1.34O2 + 1.98HCO– + 0.98H2O 0.021C5H7NO2 + 0.98NO– +
2.02H2O + 1.88H2CO3
This equation shows that the nitrification oxygen demand is 3.1 g O2 per g N.
Most process engineers use 4.2 g O2 per g N, but this overestimates the actual
nitrification oxygen demand by around 35 per cent.
The above equations show that only 0.021 mole of nitrifying bacterial cells
are produced per mole of ammonium-N nitrified, or 0.17 g of cells per g N
nitrified. This very low yield reflects the fact that nitrifying bacteria grow very
Essential Microbiology and Biology 31
slowly (see Chapter 20). The equations also show that the nitrification of 1
mole of ammonium-N consumes 1.98 moles of bicarbonate alkalinity, or 8.6 g
of HCO3 per g N nitrified. Expressing alkalinity in its usual unit of CaCO3 is
equivalent to 7.1 g CaCO3 alkalinity per g N nitrified (1 g of alkalinity as
CaCO3 = 1.22 g HCO3). If the wastewater to be nitrified does not contain this
amount of alkalinity, then alkalinity must be added (usually as sodium
bicarbonate); otherwise the reaction will stop (and not restart until sufficient
alkalinity is added).
Denitrification is the reduction of nitrate to nitrogen gas. It is an anaerobic (or
at least an anoxic) reaction achieved by many species of anaerobic and
facultative heterotrophic proteobacteria, including those in the genera
Achromobacter, Alcaligenes, Micrococcus, Pseudomonas and Thiobacillus.
Nitrate may not always be reduced to N2 – a variety of nitrogen-based gases
may be produced.
The equation for denitrification with, for example, methanol as the carbon
source is (Horan, 1990):
NO– + 1.08CH3OH + 0.24H2CO3 0.06C5H7NO2 + 0.47N2 +
1.68H2O + HCO3
This equation shows not only that oxygen is not required for denitrification,
but also (and more importantly) that bicarbonate alkalinity is generated as a
result of denitrification – nearly half the alkalinity consumed by nitrification is
regenerated by denitrification; this is important in combined
nitrification–denitrification systems (as now practised, for example, in the
‘enhanced pond systems’ in Melbourne, Australia – Chapter 9).
Photosynthetic bacteria
Purple and green anaerobic phototrophs are found in facultative ponds (see
‘Purple ponds’ in Chapter 11). In this habitat they are important as they
oxidize sulphides entering the facultative pond from the preceding anaerobic
pond (or sewer), and they thus protect us from odour and the pond algae from
the toxic effects of sulphides. They do not produce oxygen during
photosynthesis (as do algae – see below and Chapter 11) as they oxidize
sulphides to sulphur, rather than water to oxygen.
Anaerobic digestion
Anaerobic digestion is a very important process in wastewater treatment in
warm climates. It occurs in anaerobic ponds (Chapter 10) and UASBs (Chapter
18) as the major process in conjunction with sedimentation, and it also occurs in
primary facultative ponds (Chapter 11) and constructed wetlands (Chapter 17).
Anaerobic digestion is achieved by obligately anaerobic bacteria and it is
essentially the conversion, under anaerobic conditions, of settled wastewater
32 Domestic Wastewater Treatment in Developing Countries
solids to ‘biogas’ – that is methane and carbon dioxide. Biogas is a valuable fuel
which, at large plants (eg the modern waste stabilization ponds at Melbourne,
Australia – Chapter 9), can be profitably recovered to generate electricity.
Anaerobic digestion proceeds in four stages:
Hydrolysis: The hydrolysis of complex wastewater organics (such as
proteins, polysaccharides and fats);
Acidogenesis: the anaerobic oxidation of fatty acids and alcohols and the
fermentation of amino acids and carbohydrates to volatile fatty acids (eg
butyrates and propionates) and hydrogen gas;
Acetogenesis: the conversion of butyrate and propionates to acetates; and
Methanogenesis: the conversions of acetates, and hydrogen and carbon
dioxide, to methane.
Many anaerobic and facultative bacterial species are responsible for Stage 1,
such as Bacillus, Clostridium, Proteus, Micrococcus, Staphylococcus and
The acidogens, responsible for Stage 2, include Butyrovibrio, Clostridium
and Eubacterium; they convert sugars to volatile fatty acids:
C6H12O6 (glucose) + 2H2O 2CH3OOH (acetic acid) + 4H2
When the concentration of H2 becomes high glucose is converted into
propionic and butyric acids:
C6H12O6 + 2H2O 2CH3CH2COOH (propionic acid) + 2CO2 + 2H2
C6H12O6 CH3(CH2)2COOH (butyric acid) + 2CO2 + 4H2
These acids are converted in Stage 3 to acetic acid by, for example,
Synthobacter and Synthrophomonas:
CH3(CH2)2COOH + 2H2O 2CH3COOH + 2H2
Stage 4 is the conversion of acetates to methane and carbon dioxide, and of
hydrogen and carbon dioxide to methane:
CO2 + 4H2 CH4 + 2H2O
The methanogens are all Archaea of, for example, the genera Methanothrix,
Methanosarcina and Methanococcus. They are very slow growing, with
generation times of ~24 hours – much longer than the bacterial groups
Essential Microbiology and Biology 33
involved in Stages 1–3 – and methanogenesis is thus the rate-limiting stage in
anaerobic digestion. The methanogens are also more sensitive than the other
groups to environmental stress (eg too low a pH) and, in high-sulphate
wastewaters, they are outcompeted by sulphate-reducing bacteria, such as
Desulfovibrio spp, for hydrogen and acetate (which the sulphate-reducers use
in their reduction of sulphate to sulphides – Chapter 10).
Overall the anaerobic treatment of domestic wastewater in warm climates
is extremely advantageous, and high removals of BOD (70–80 per cent) at
short retention times (8–24 hours) are achieved in anaerobic ponds and
Faecal indicator bacteria
The concept of faecal indicator bacteria was developed in the late 19th century
to assess the efficacy of water treatment: if bacteria of exclusively faecal origin
are found in a water, then we know that the water has been polluted by faeces
and that it may, therefore, contain pathogenic faecal bacteria (ie those in
Category II, Chapter 2). Conversely, if treated drinking water is shown not to
contain any faecal indicator bacteria, then it is unlikely to contain any
pathogenic micro-organisms. Of course, with wastewaters the situation is
different: we know that wastewaters are faecally polluted – they contain faeces
and faecal micro-organisms, including faecal pathogens of most, if not all, of
Categories I–V. We use the numbers of faecal indicator bacteria in wastewaters,
therefore, not to indicate faecal pollution, but to indicate faecal pathogen
removals in wastewater treatment processes, and to estimate the health risks
in wastewater re-use (Chapters 21 and 22). This works very well for faecal
bacterial pathogens, quite well for faecal viral pathogens, but not at all well
for faecal protozoan and helminthic pathogens.
The requirements for an ‘ideal’ faecal indicator bacterium, as applied to
wastewater (rather than drinking water), are that:
it should be exclusively faecal in origin,
its numbers in wastewater should be greater than those of faecal viral and
bacterial pathogens,
its removal in wastewater treatment processes should be close to that of
faecal viral and bacterial pathogens, and
it should be simple and inexpensive to count its numbers reliably and
As might be expected, no bacterium always meets all these requirements, but
one comes very close: one (and only one) of the coliform bacteria, namely
Escherichia coli.
Coliform bacteria
The early water bacteriologists identified the coliform group of bacteria as
faecal indicator organisms. This group was considered in two parts: ‘total
34 Domestic Wastewater Treatment in Developing Countries
coliforms’ and ‘faecal coliforms’, with the former comprising both non-faecal
and faecal coliforms. It was originally considered in this way because counting
faecal coliforms was initially a two-stage process: first the number of total
coliforms in a drinking water sample was determined and then, if any of these
were present, the number of faecal coliforms was determined (many water
bacteriologists still use the same basic procedure today). However, with
wastewater, it is only the faecal coliforms that are relevant (it is basically
meaningless to report total coliform numbers in wastewaters, especially
tropical and subtropical wastewaters).
Almost all coliforms, faecal and non-faecal alike, oxidize the disaccharide
lactose with the production of acid and gas at 37°C (human body temperature)
in the presence of bile salts (which are used to inhibit non-intestinal bacteria,
although it is now more common to use a surface-active agent, such as Triton
X100, with similar growth-inhibiting properties). Lactose comprises equal
proportions of the two monosaccharides glucose and galactose, and only
coliform bacteria possess the enzyme β-galactosidase which enables it to break
down galactose. Thus the modern definition of a coliform bacterium is one
that possesses β-galactosidase (or, strictly speaking, the gene that codes for this
enzyme). Only faecal coliforms can break down lactose to acid and gas at the
higher temperature of 44°C. However, this is true only in temperate regions
(and even there not always true); in tropical and subtropical regions some nonfaecal coliforms can produce acid and gas from lactose at 44°C, and some true
faecal coliforms may not produce any gas from lactose (ie they are
‘anaerogenic’). Thus the concept of faecal coliforms is not strictly applicable in
warm climates, although the majority of acid and gas producers at 44°C are
indeed faecal in origin. Even so, it is now considered much better to count
non-faecal coliforms (due to the small proportions of false positives and false
negatives – that is non-faecal coliforms growing at 44°C, and some faecal
coliforms either not able to grow at this elevated temperature or unable to
produce gas from lactose, respectively), but rather to count the single coliform
bacterium that really is exclusively faecal in origin. This bacterium is
Escherichia coli.
Escherichia coli
The early water bacteriologists counted (or, more correctly, tried to count)
the bacterium then known as Bacterium coli communis (Smith, 1895 – this
is the earliest reference to a faecal indicator bacterium). In fact, it was the
difficulty of counting only this bacterium that led, in the early 20th century,
to the use of total and faecal coliforms to assess the quality of drinking
waters, as these could be counted with at least reasonable reliability.
Bacterium coli communis is now called Escherichia coli. Like all coliforms,
E coli has the enzyme β-galactosidase but, uniquely amongst the coliforms, it
also has the enzyme β-glucuronidase (which it uses to break down
glucuronate to glucose and uronic acid). Modern media to detect or count E
coli contain a chromogenic substrate to detect this enzyme and thus give the
resulting colonies of E coli a distinct colour (usually blue or purple); several
Essential Microbiology and Biology 35
such media are commercially available (eg ‘Chromagar E coli’, on or in which
E coli forms blue colonies when incubated at 37°C for 24 h; Chromagar,
2002*). Given the ease with which specifically E coli counts can be obtained,
it is now time (especially in warm climates, but also in temperate climates) to
cease counting faecal coliforms and to determine only the numbers of E coli.
This removes the problem of faecal coliform counts including some nonfaecal coliforms since E coli is an exclusively faecal micro-organism (and the
only faecal coliform currently so recognized – see Edberg et al, 2000*;
Leclerc et al, 2001*).
In this book references to faecal coliforms in standards and guidelines (eg
those of the Council of the European Communities and the World Health
Organization) have been replaced by references to E coli. This permits a better
interpretation of the intent of these standards and guidelines (and it is to be
hoped that future revisions of them will use E coli rather than faecal
coliforms). However, when reported results of faecal coliform numbers
(obtained experimentally or from monitoring programmes) are referred to,
this change is not made and the reported designations of faecal coliforms and
‘FC numbers’ are retained.
Protozoa are single-celled eukaryotes. A few are important human pathogens
– Giardia, Cryptosporidium, Cyclospora and Entamoeba, for example, are
major excreta-related pathogens which are consequently present in domestic
wastewaters (Category I, Chapter 2). (The genus Plasmodium contains the
malaria parasites, but malaria is a water-related, rather than an excreta-related,
disease.) However, most protozoa are non-pathogenic and very widely
distributed in nature.
The protozoa can be conveniently classified into three groups: amoebae,
ciliates and flagellates. The last two groups are important in wastewater
treatment: flagellates in the class Zoomastigophora are present in very large
numbers in wastewater treatment processes, as are many species of ciliates.
Flagellate biomass is generally higher than that of the ciliates, although there
is a greater species diversity of the latter.
Flagellates generally grow heterotrophically; in wastewater treatment
reactors they are thus in competition with the more efficient bacterial
heterotrophs. The ciliates display a wider range of morphology and nutrition:
some are ‘free-swimming’, others are ‘crawling’ organisms, and yet others have
a stalk which attaches to particulate material (such as an activated sludge floc)
(Figure 3.5). Ciliate nutrition is mainly by ‘phagocytosis’ – that is they engulf
other microbes (bacteria, algae and other protozoa) and digest them
enzymatically (an early form of ‘eating’ as we know it).
Protozoa have been extensively studied in conventional wastewater
treatment processes such as activated sludge and biofilters (Chapters 19 and
20). A healthy protozoan population in activated sludge aeration tanks
36 Domestic Wastewater Treatment in Developing Countries
Source: Water Pollution Control Laboratory (1968)
Figure 3.5 Five of the Commonest Ciliated Protozoa Found in Wastewater
Treatment Works: (a) Chilodonella uncinata, (b) Opercula microdiscum,
(c) Aspidisca costata, (d) Trachlophyllum pusillum and (e) Carchesium
polypinum; (a) and (c) are crawling ciliates, (b) and (e) stalked ciliates and
(d) a free-swimming ciliate
significantly improves the treatment efficiency – effluent suspended solids
concentrations are ~70 per cent less with ciliated protozoa than without them.
Ciliated protozoa have recently been shown to be responsible for some of
the removal of Cryptosporidium oocysts in constructed wetlands (Chapter 17)
Essential Microbiology and Biology 37
(Stott et al, 2001*, 2003*). Laboratory tests showed ingestion rates of >50
oocysts per hour by Paramecium, although isolates from constructed wetlands
ingested 4–10 oocysts per hour. The role of ciliated protozoa in waste
stabilization ponds (Chapters 9–12) has yet to be determined.
The micro-algae in facultative and maturation waste stabilization ponds
(Chapters 11 and 12) are single-celled Eukarya. The cells are green as, like the
leaves of higher plants, they contain large amounts of chlorophyll, the pigment
that captures light energy in photosynthesis. The algae use this energy to fix
carbon dioxide which is their main source of carbon (so algae are photoautotrophs), although they can grow photoheterotrophically on simple organic
compounds (such as acetate). During photosynthesis oxygen is produced from
water, and in facultative and maturation ponds this is the main source of
oxygen used by the bacterial heterotrophs in the ponds for the removal of
BOD. The algae, when they are photosynthesizing rapidly, induce a high pH in
the ponds (especially in maturation ponds); the pH can rise to >9.4, which is
critical for faecal bacterial die-off in ponds. Further details are given in
Chapters 11 and 12.
Helminths (worms) are important because a few of them cause disease
(Categories III–V, Chapter 2) and because a group of them are highly tolerant
of pollution and oxygen depletion in freshwaters (see below). All worms fall
into one of three types: nematodes (roundworms), cestodes (flatworms) and
trematodes (flukes). Many, especially those that are pathogenic, have quite
complicated life cycles, often both in and outwith the human body, but despite
this they are extremely successful human parasites: around one-third of the
world’s population (ie ~2 billion people) is infected with one or more worms.
As a result, worm egg numbers in wastewaters in developing countries (in
which almost all the worm infections occur) are generally high – up to a few
thousand/l in newly sewered communities, although over time (as the
opportunities for reinfection decrease as a result of more and more
communities being sewered) the numbers decline to <1000 and eventually to
<100 or even <10/l.
Egg numbers in treated wastewaters must be reduced to very low levels
when the treated wastewaters are used for crop irrigation and/or fish culture
(Chapters 4, 21 and 22). For crop irrigation the number of human intestinal
nematode eggs (ie those of the geohelminths – Category III, Chapter 2) should
be ≤1/l or, if children under the age of 15 are exposed, ≤0.1/l. For fish culture
trematode eggs (ie those of the water-based helminths – Category V) must be
absent as these worms multiply tens of thousands of times in their first
38 Domestic Wastewater Treatment in Developing Countries
intermediate aquatic host (an aquatic snail), and thus the eggs from one person
can potentially infect many hundreds of people.
Well illustrated on-line reference texts include Muller (2001*) for human
worm diseases and Ayres and Mara (1996*) for counting the numbers of
human intestinal nematode eggs in wastewaters.
Invertebrates are animals without backbones and in clean unpolluted streams,
rivers and lakes there are many different types of small invertebrates. These may
be present as adult and juvenile forms, or only the latter (such as larvae and
nymphs). Some of these ‘micro-invertebrates’ are very sensitive to aquatic
pollutants, and some are very tolerant of pollution. We can therefore use the
micro-invertebrate fauna of a freshwater to assess its ‘health’ – that is to
determine biologically the extent of pollution. The number of different microinvertebrates present can be used to develop a ‘biotic index’ of freshwater quality
(and this is used to complement physicochemical water quality – Chapter 4).
A large amount of work has been done on assessing freshwater quality in
this way, mainly in industrialized countries (eg Welch, 1992; Wright et al,
2000; Adams, 2002; Greenwood-Smith, 2002), with much less application in
developing countries (Madhou, 2000; see also Girgin et al, 2003*). In many
cases there is very little or no information on micro-invertebrates in unpolluted
waters in developing countries; and, even though many waterbodies are
seriously polluted, there is not much information on the biological quality of
polluted waters. Dudgeon (2002*) details the adverse impacts of human
activities on aquatic biodiversity in monsoonal Asia: pollution, excessive
exploitation (ie overharvesting), and conversion of riverine wetlands to
agriculture have led to dramatic decreases in biodiversity – fewer riverine birds,
endangered turtle populations, declining fish populations, and so on.
Biotic index for tropical freshwaters
A simplified biotic index for tropical and subtropical freshwaters is presented
by Van Damme (2001*). Ten groups of micro-invertebrates are used, and these
are placed into three categories based on their sensitivity to, or tolerance of,
pollution (Figure 3.6 and Table 3.1). The presence and relative abundance of
members of these groups in a water sample define an approximate biotic index
(on a scale of 0–10, with 0 indicating gross pollution and 10 indicating
excellent water quality) (Table 3.2).
Procedures for sampling and analysis (ie micro-invertebrate identification
using a stereoscopic microscope) are given by Van Damme (2001*) and also in
Standard Methods (American Public Health Association, 1998). The
procedures are very simple and require only minimal equipment – indeed
obtaining biotic indices as described by Van Damme is a very good biology
class project for secondary school students and university engineering students
(see De Pauw et al, 1999).
Essential Microbiology and Biology 39
Indicator Group I: Macroinvertebrate groups highly sensitive to oxygen depletion/pollution
Freshwater mussels
Damselfly nymphs
Indicator Group II: Macroinvertebrate groups moderately sensitive to oxygen depletion/pollut
Water beetles
Dragonfly nymphs
Water bugs
Freshwater shrimps
Hog lice
Freshwater snails
Freshwater crabs
Indicator Group III: Macroinvertebrate groups little sensitive to oxygen depletion/pollution
Mosquito larvae
Hover-fly maggots
Oligochaete worms
Source: Van Damme (2001*)
Figure 3.6 Micro-invertebrates Used to Assess the Biological Quality of
Tropical Waters
40 Domestic Wastewater Treatment in Developing Countries
Table 3.1 Micro-invertebrate Groups Used to Assess the
Biological Quality of Tropical Waters
Indicator group
Sensitivity to
oxygen depletion
Damselfly nymphs
Freshwater mussels
Water beetles, water bugs, hog lice
Dragonfly nymphs
Freshwater snails
Shrimps and crabs
Leeches, flatworms
Mosquito larvae
Hoverfly maggots
Oligochaete worms
Source: Van Damme (2001*)
Table 3.2 Simplified Biotic Index for Tropical Waters
Biotic index
Water quality
Very bad
Note: a See Table 3.1
Source: Van Damme (2001*)
Micro-invertebrate groups presenta
Many Group I
Some Group I
No Group I, many Group II
Some Group II, some Group III
Only Group III
Effluent Quality
Wastewater treatment must be done for a specified purpose – for example, to
produce an effluent suitable for agricultural or aquacultural reuse (or both), or
to produce an effluent that can be safely discharged into inland or coastal
waters. Wastewater treatment plant designers have to know what is going to
happen to the effluent – re-use or discharge – before they design the plant, as
the effluent quality requirements will vary accordingly.
Effluent quality requirements, often termed effluent quality standards, are
set by regulatory agencies that are empowered by legislation to make such
regulations – environmental protection agencies, for example. The regulations
that these agencies make are legally binding on the authority responsible for
wastewater treatment. Regulatory agencies have a duty, either explicitly
defined in the governing legislation or at any rate implicitly, to set sensible
regulations. Unfortunately, in many developing countries not all such
regulations are as sensible as they should be (Johnstone and Horan, 1994,
1996*), and this also occurs in industrialized countries (Dolan, 1995; Mara,
1996*). Wastewater treatment engineers need to understand how effluent
quality standards should be set properly so that, if necessary, they can have a
rational dialogue with the regulatory agency to ensure that it does not impose
standards that are too high. This is extremely important since compliance with
standards that are too high requires an unnecessary expenditure of money
(generally on inappropriate high-tech wastewater treatment systems – see von
Sperling and Chernicharo, 2002*), and the people who ultimately pay this
unnecessarily large amount of money are the local people who generate the
wastewater that is treated to too high a standard. Incomes fall, and falling
incomes mean poorer health, even deaths. In the US, for example, Miller and
Conko (2001*) report that one death is estimated to result from every US$7.25
million spent on regulatory costs through this income effect. As these authors
note, ‘the expression “regulatory overkill” is not merely a figure of speech’.
If a regulator insists on a very strict national standard for one or more
parameters prior to river discharge, wastewater treatment may be so expensive
that no city can afford to treat its wastewater. From the regulator’s perspective
this is a self-defeating situation: treatment plants are not built and the regulator
42 Domestic Wastewater Treatment in Developing Countries
has no effluents to regulate. As regulators are meant to protect the
environment, this is a truly nonsensical situation: no treatment plants means
continued discharges of untreated wastewater, and hence unabated freshwater
pollution, continued risks to human health, and no environmental protection
Using treated wastewaters for crop irrigation or for fishpond fertilization is a
very sensible thing to do, especially in water-short areas. However, it must not
cause any excess transmission of excreta-related disease, and therefore the
wastewaters must be treated to an appropriate microbiological quality. A
detailed discussion on what constitutes an appropriate microbiological quality
is given in Chapter 21 for agricultural re-use and in Chapter 22 for
aquacultural re-use. Here only the basic principles are addressed.
The re-use of untreated wastewaters in agriculture or aquaculture is known
to cause an excess transmission of certain excreta-related diseases, especially
those in Categories II, III and IV – the bacterial and geohelminthic diseases,
and the water-based trematode diseases (Chapter 2). Thus treatment to remove
faecal bacterial pathogens and human intestinal nematode and trematode eggs
from the wastewater is essential – but removal to what degree? The answer to
this question is that they must be removed to a level which does not cause
excess disease in the people working in the wastewater-irrigated fields or
wastewater-fertilized aquaculture ponds, or in those who consume the
wastewater-irrigated crops or wastewater-fertilized aquacultural produce (fish,
for example). The next question is obvious: what is the level that does not
cause excess infection? Expressed another way, this question means: what are
the safe minimum microbiological requirements for wastewater treatment?
The answer to this question is certainly not the United States Environmental
Protection Agency (USEPA)/United States Agency for International
Development (USAID) requirement that, for wastewater-irrigated salad crops
and vegetables eaten uncooked, treated wastewaters should contain zero E coli
per 100 ml (Environmental Protection Agency, 1992), since this is the
bacteriological requirement for drinking water and therefore a good example
of regulatory overkill.
The World Health Organization has produced guidelines for the
microbiological quality of treated wastewaters used in agriculture and
aquaculture (WHO, 1989*, 2004*; see also Blumenthal et al, 2000*). These
for ‘restricted’ irrigation – that is the irrigation of all crops except salad
crops and vegetables eaten uncooked:
≤1 human intestinal nematode eggs/l (reduced to ≤0.1 eggs/l when children
under 15 years are exposed), and
≤105 E coli/100 ml.
Effluent Quality 43
The nematodes are Ascaris, Trichuris and the human hookworms (Chapter 2).
for ‘unrestricted’ irrigation – that is including salad crops and vegetables
eaten raw:
≤1 human intestinal nematode eggs/l (also reduced to ≤0.1 eggs/l when
children under 15 years are exposed through their fieldworker-parents
bringing home ‘unrestricted’ produce directly from the fields), and
≤1000 E coli/100 ml.
The epidemiological and experimental evidence for these guidelines is reviewed
in Chapter 21, which also gives risk calculations, based on quantitative
microbial risk analysis procedures, to support the E coli guideline of
≤1000/100 ml for unrestricted irrigation.
for aquacultural re-use:
zero viable trematode eggs/l of treated wastewater, and
≤1000 E coli/100 ml of aquaculture pond water.
The trematodes of importance are Schistosoma spp, Clonorchis sinensis and
Fasciolopsis buski. The rationale for these guidelines is given in Chapter 22.
The risks to public health when these guidelines are applied are extremely
small, and certainly less than the World Health Organization’s (2003*)
tolerable risk of infection of 10–3 per person per year – that is it is considered
acceptable if one person in every 1000 becomes infected in a 12-month period
from consuming salad crops and vegetables irrigated, or fish from fishponds
fertilized, with wastewater treated to these guideline levels (see Chapter 21).
In addition to not harming human health, treated wastewaters should not
harm the crops, and thus they should meet the physicochemical quality
requirements for all waters used for irrigation given by the Food and
Agriculture Organization (Ayers and Westcot, 1989*). For treated domestic
wastewaters, the most important of these are: electrical conductivity, sodium
absorption ratio, boron concentration, total nitrogen concentration and pH.
The precise values of these parameters depend on the types of crops being
grown as different crops have different sensitivities to them. Details are given
in Chapter 21.
If a treated wastewater is discharged into a river it exerts a demand on the
oxygen resources of the river. This removal of dissolved oxygen (DO) for
wastewater oxidation must be balanced by an addition of oxygen. The most
important source of oxygen for reoxygenation of the river is the atmosphere:
there is a mass transfer of oxygen from the atmosphere across the water surface
to the river water below. The rate of this transfer is proportional to the oxygen
deficit in the water (ie the difference between the oxygen saturation
44 Domestic Wastewater Treatment in Developing Countries
DO deficit (mg/l)
Flow time (d)
Note: The oxygen sag curve is the sum of the deoxygenation and reoxygenation processes in the
river (bacterial oxidation and surface reaeration, respectively). The wastewater is added to the
river flow at zero time; the initial deficit D0 decreases to a maximum or critical deficit Dc at a
distance downstream often equivalent to several days flow time.
Figure 4.1 The Dissolved Oxygen Sag Curve
Effluent Quality 45
concentration and the actual DO concentration). Thus the DO removal that
occurs below the point of discharge actually stimulates an increased rate of
supply of oxygen from the atmosphere. This competition between
deoxygenation and reoxygenation results in a DO profile which typically
shows a distinct ‘sag’ some distance below the point of discharge (Figure 4.1).
In order to prevent the river becoming anaerobic, there must be an adequate
DO reserve at all points along the river. Analysis of the oxygen sag curve
provides a convenient method of determining the degree of treatment that
should be given to the effluent before it is discharged, so as to ensure that the
lowest DO concentration that occurs is not less than the minimum required to
maintain the river water quality at the desired level.
DO sag curve analysis
The first analysis of the DO sag curve was made by Streeter and Phelps (1925)
for the Ohio River in the US; full details are given in Phelps (1944). For a single
wastewater effluent discharged into a long river with a reasonably constant
flow regime, the DO sag curve results from the competition between DO
demand and DO supply. The former is due to the ultimate BOD (biochemical
oxygen demand) of the wastewater effluent, and the latter comes primarily
from reoxygenation through surface reaeration by the oxygen in the
The DO deficit (D, mg/l) is defined as the difference between the oxygen
saturation concentration (ie its solubility) in the river water at the river water
temperature and the actual DO concentration in the river water. The DO
demand due to the ultimate BOD of the effluent increases the deficit, and the
DO supply by reaeration decreases it. Thus, using the general format of
equation 5.1, the Streeter–Phelps equation is:
dD/dt = k1L – k2D
where D is the DO deficit at time of flow t (ie assuming a constant velocity of
flow in the river, at a given distance downstream of the effluent discharge
point), mg/l; k1 is the first-order rate constant (base e) for BOD removal in the
river, day–1; L is the ultimate BOD of the effluent–river-water mixture at time
t, mg/l; and k2 is the first-order rate constant (base e) for surface reaeration,
Equation 4.1 can be integrated to yield:
–k t
–e 2
+ D0 e–k2t
where D0 and L0 are the values of D and L at t = 0 (ie at the point of
46 Domestic Wastewater Treatment in Developing Countries
The most important point in the sag curve is the maximum DO deficit,
called the ‘critical’ deficit, Dc (Figure 4.1). At this point in the curve dD/dt is
zero, so from equation 4.1:
Dc = k1L/k2
The time (or distance) at which Dc occurs is tc, given by substituting equation
4.3 into equation 4.2:
tc =
D0 (k2–k1)
The value of k2 can be estimated from the following equations (Melching and
Flores, 1999*):
‘Pool and riffle’ streams with a flow Q < 0.556 m3/second:
k2 = 517(vS)0.574Q–0.242
Pool and riffle streams with Q > 0.556 m3/s:
k2 = 596(vS)0.528Q–0.136
‘Channel-control’ streams with Q < 0.556 m3/s:
k2 = 88(vS)0.313H–0.353
Channel-control streams with Q > 0.556 m3/s:
k2 = 142(vS)0.333H–0.66W–0.243
where v is the average streamflow velocity in the length of stream considered,
m/s (range used for the derivation of equations 4.5–4.8: 0.003–1.83 m3/s); S is
the slope of the water surface, m/m (range: 0.00001–0.06 m/m); Q is the
stream flow, m3/s (range: 0.0028–210 m3/s); H is the average stream depth
calculated from the continuity equation H = Q/(vW), m (range: 0.0457–3.05
m); and W is the average stream top width, m (range: 0.78–162 m). Equations
4.5–4.8 were derived from observations on 500 reaches in 166 streams and
rivers throughout the US.
Effluent Quality 47
A ‘pool and riffle’ stream alternates between pools (deep areas, with a
nearly horizontal water surface at low flows) and riffles (shallow, high-velocity
areas); bed material is sand and gravel in the size range 2–246 mm. ‘Channelcontrol’ streams are characterized by reasonably uniform, steady flow with
width-to-depth ratios often >40 and water-surface slopes of <0.04; the
‘control’ can be achieved by hydraulic structures (eg weirs or dams).
In the case of treated wastewater discharges to lakes, k2 values depend on
wind speed (but not on fetch, which is the horizontal wind–water-surface
contact distance), as follows (Gelda et al, 1996*):
k2 = α(U10)β/H
where α is a constant determined by model calibration; U10 is the wind speed
at a height of 10 m above the lake surface, m/s; β is a constant (taken as 1 for
U10 ≤ 3.5 m/s and 2 for U10 > 3.5 m/s); and H is the mean lake depth, m.
Limitations of the Streeter–Phelps equation
The simple Streeter-Phelps equation given above considers only a single point
discharge, one oxygen sink (the BOD of the river–effluent mixture), and one
oxygen source (surface reaeration). In practice there may be more than one
discharge, or the river may receive diffuse pollution from, for example,
agricultural run-off. Additionally, plants in the river or growing on its banks
may supply oxygen by photosynthesis. However, the most important
additional factor to consider, especially in rivers in developing countries, is the
so-called ‘benthic’ oxygen demand. This is the amount of oxygen used by the
bacteria in the bottom mud of the river which, due to earlier discharges of
untreated wastewater, can often be very high. Dobbins (1964) took this into
account, and his version of the Streeter–Phelps equation modified for benthic
demand is:
dD/dt = (k1 + k3)L – k2D
where k3 is the first-order rate constant for oxygen consumption (ie BOD
removal) in the bottom mud, d–1.
Dobbins also developed an equation for D which took into account the
other factors mentioned above, but it is really too complicated to use – it is
now much better to use one of the commercially available computer models
for river water quality, for example, MIKE 11 (DHI Software, 2002*) or River
Water Quality Model 1 (Shanahan et al, 2001*, IWA Task Group, 2001).
The above, essentially introductory, discussion on dissolved oxygen supply
and demand in a river subject to effluent discharges assumes first order
kinetics. In practice this assumption is satisfactory for the in-river processes of
DO supply and demand, but much less satisfactory for the determination of
the ultimate BOD of the effluent (ie as determined by equation 5.4). This point
48 Domestic Wastewater Treatment in Developing Countries
is considered in more detail in Chapter 5, but it should be noted that an error
of underestimation of the ultimate BOD of an effluent results in a
corresponding underestimation in the oxygen demand of the river–effluent
mixture, and thus introduces error into estimates of the dissolved oxygen
balance in the river downstream of the effluent discharge point.
Index of physicochemical river water quality
Regulators generally classify rivers to assess their water use for various
purposes (eg potable supply or irrigation) on the basis of their physicochemical
quality. An early attempt at this, based on a single parameter, is given in Table
4.2 below. A more comprehensive river water quality index is used by the
Environment Agency (for England and Wales): it is based on eight parameters
(dissolved oxygen, nitrification-suppressed BOD, total and unionized
ammonia, pH, hardness, dissolved copper and total zinc), with maximum
values specified for each parameter for each of five classes of river water
quality, which range from RE1 (highest quality) to RE5 (lowest). The
maximum values are specified on a percentile basis (eg a 95-percentile
requirement means that only 5 per cent of samples are allowed to be above the
specified maximum). Full details are given in Martin (2002*) and UK
Legislation (1994*).
Of course, river water quality indices and classifications developed in
industrialized countries are not directly applicable to rivers in developing
countries, but they may serve as a general guide. Madhou (2002), for example,
has proposed a river water quality index for Mauritius by adapting
industrialized country indices for local conditions.
‘Minimal’ water quality index
Working on the Suquía River in central Argentina, Pesce and Wunderlin
(2000*) proposed a very simple river water quality index, which they
designated WQImin. It is based on only three parameters: dissolved oxygen,
mg/l; total dissolved solids (ie salts in solution), mg/l; and turbidity, NTU (ie
turbidity units, dimensionless). WQImin, which has a score range of 0–100
(grossly polluted to extremely high quality), is defined as:
WQImin = CDO + CTDS + CTurb
where CDO, CTDS and CTurb are the ‘normalized unit values’ for dissolved
oxygen, total dissolved salts and turbidity, respectively, as given in Table 4.1
for ranges of individual measured parameter concentrations. WQImin was
found to correlate well with more comprehensive indices which included 20
parameters, although it was recommended that one of the more complete
indices be evaluated a few times each year, with WQImin being used for routine
analysis (once or twice per month).
Effluent Quality 49
Table 4.1 Normalized Unit Values for Dissolved Oxygen, Total Dissolved
Salts and Turbidity Used to Calculate WQImin
DO Concentration
≥ 7.5
≥ 1.0
Total dissolved salts
concentration (mg/l)
≤ 20,000
Turbidity (NTU)
≤ 100
Normalized unit
value (C)
Example: Suppose the DO concentration was measured as 6.8 mg/l, the Total dissolved salts
concentration as 790 mg/l and the turbidity as 16 NTU. Thus CDO = 80, CTDS = 70 and CTurb =
70; and WQImin = (80 + 70 + 70)/3 = 73.
Source: Pesce and Wunderlin (2000*)
Effluent standards
It is administratively more convenient to enforce an effluent standard rather
than a stream or river standard. The local regulatory agency should select
quality standards for wastewater effluents which ensure that rivers do not
become unsuitable for their present use or intended purpose. The best known
and most widely (and usually inappropriately) applied effluent standard is the
so-called ‘20/30 Royal Commission standard’ (ie ≤20 mg BOD5/l and ≤30 mg
SS/l). The United Kingdom Royal Commission on Sewage Disposal of
1898–1915 was appointed to consider appropriate methods of sewage
treatment and disposal in the United Kingdom (it is necessary to stress ‘the
United Kingdom’ because the Commissioners’ recommendations were meant
to apply only to this country, although they are often indiscriminately applied
to other countries in different climatic zones – see Johnstone and Norton,
2000). The Commissioners classified British rivers on the basis of their 65 °F
(ie 18.3°C) BOD5, as shown in Table 4.2. The Commissioners chose a 65 °F
BOD because the long-term average summer temperature in the UK is 65 °F,
and they chose 5 days as this gave the most reliable and consistent BOD test
results. (The standard BOD5 test is now conducted at 20°C rather than
If an effluent is discharged into a river, a mass balance of BOD at the point
of discharge (Figure 4.2) yields:
LrQr + LeQe = Lm(Qr + Qe)
50 Domestic Wastewater Treatment in Developing Countries
Table 4.2 The UK Royal Commission’s Classification of River Water Quality
River classification
BOD5 (mg/l)
Very clean
Fairly clean
where L is BOD5, mg/l (= g/m3); Q is flow, m3/day; the subscript r refers to the
river just upstream of the point of discharge; e refers to the effluent; and m
refers to the river–effluent mixture just downstream of the point of discharge.
Now if the effluent is diluted with eight volumes of clean river water (ie if
Qr/Qe = 8 and Lr = 2 mg/l), the maximum BOD5 of the effluent to avoid
nuisance in the river (ie for Lm = 4 mg/l) is given by equation 4.12 as:
Le = [Lm(Qr + Qe) – LrQr]/Qe
= Lm[Qr/Qe) + 1] – Lr(Qr/Qe)
= 4(8 + 1) – (2 – 8)
= 20 mg/l
To this standard for BOD5 the Commissioners added their standard of
≤30 mg/l for suspended solids (SS). Although the 20/30 standard is usually
referred to as the Royal Commission standard, the Commissioners in fact
recommended standards for each of five ranges of available dilution
Lm, Qr, Qe
Figure 4.2 Discharge of an Effluent into a River – definition of terms used in
the BOD mass balance
Effluent Quality 51
Table 4.3 The UK Royal Commission’s Standards for Wastewater Effluents
Discharged into Rivers
Available dilution (volumes of clean river water per unit
volume of effluent)
Maximum permissible
concentration (mg/l)
a No standard recommended; theoretically infinite
b Exact values to be decided on the basis of local circumstances
(Table 4.3). There is no evidence to suggest that these Royal Commission
standards are directly applicable to climates other than those similar to that in
the UK. In many developing countries the ‘natural’ BOD of a river can be high
– for example, the unpolluted River Turkwell in the remote area of northern
Kenya has a BOD5 of 20–45 mg/l (Meadows, 1973).
In many developing countries effluent standards do not exist. Even so
design engineers need to ensure that the effluent produced in their treatment
works will not pollute the receiving watercourse. A standard is certainly
required for BOD (or COD – chemical oxygen demand) and possibly for
suspended solids in the case of discharge to inland and coastal waters, for E
coli in the case of coastal discharges, and for E coli and nematode or trematode
worm eggs in the case of wastewater re-use in agriculture or aquaculture.
BOD standard
Often a national ‘blanket’ BOD standard is specified, for example, ≤30 mg/l in
India (Central Pollution Control Board, 1996) and ≤25 mg/l in the European
Union (Council of the European Communities, 1991a*). This appears to
follow the assumed logic of the UK Royal Commission’s BOD ‘standard’ of
≤20 mg/l, but ignores the reasoning behind this value (dilution with eight
volumes of clean river water), the available dilution in the receiving river (Table
4.3), and any river water quality objective (eg a required minimum DO level).
A useful document for regulators is the report by the former National Rivers
Authority (for England and Wales, now the Environment Agency) which sets
out its rationale for establishing effluent quality requirements (National Rivers
Authority, 1994; see also Martin, 2002*).
Waste stabilization ponds (WSP) in the European Union are required to
comply with the above 25 mg/l standard for BOD, but on a filtered basis –
that is samples are to be filtered before analysis to remove the BOD due to the
algae present. This is discussed in more detail in Chapter 11.
In general, a BOD standard should be interpreted as a carbonaceous BOD
standard (CBOD) excluding, therefore, the BOD due to the oxygen demand of
52 Domestic Wastewater Treatment in Developing Countries
nitrification (Chapter 3). The addition of 2-chloro-6-(trichoromethyl)pyridine
to a final concentration of 10 mg/l in the BOD test bottles inhibits nitrification
(American Public Health Association, 1998). However, CBOD should not be
used for raw or settled (eg anaerobic pond effluent – Chapter 10) wastewater
as it underestimates their strength by as much as 20–40 per cent, and can
therefore result in an underdesign of the same amount (Albertson, 1995).
Suspended solids
In India the suspended solids standard is ≤100 mg/l; and in the European
Union ≤35 mg/l, although this is only an ‘optional’ requirement, but WSP
effluents have to contain ≤150 mg/l. Many rivers in developing countries have
a much higher ‘natural’ suspended solids concentration, and there is little point
in setting a standard more stringent than the background SS concentration in
the receiving water.
Ammonia in wastewater is a mixture of dissolved ammonia gas (NH3) and
the ammonium ion (NH4), often referred to as ‘free’ and ‘saline’ ammonia,
respectively. Free ammonia at concentrations >0.5 mg N/l is toxic to fish
(Chapter 22). In India, the standard for total (ie free and saline) ammonia is
≤50 mg N/l and for free ammonia ≤5 mg N/l. There is also a standard for total
Kjeldahl nitrogen (organic N and ammonia) of ≤100 mg/l. In the European
Union there is a total nitrogen requirement of ≤15 mg N/l for populations up
to 100,000 and ≤10 mg N/l for larger populations, but this is required only for
discharge to ‘sensitive’ areas (defined as areas which are ‘subject to
eutrophication’). Total nitrogen is the sum of organic N, ammonia, nitrite and
E coli
As discussed earlier in this chapter and also in Chapters 21 and 22, there are
specific requirements for E coli (also for nematode and trematode eggs) when
treated wastewaters are used in agriculture and aquaculture. The question to
be answered here is: should there be an E coli standard for discharges to inland
In developing countries inland surface waters are often used by rural and
periurban communities as their domestic water supply, commonly without
treatment. Studies in the Philippines have shown that young children are at a
high risk of diarrhoeal disease when their drinking water contains >1000
E coli/100 ml, but not when it contains less (Moe et al, 1991). Ideally,
therefore, river waters should contain ≤1000 E coli/100 ml. Knowing the
minimum available dilution in the receiving water and its E coli count
upstream of the discharge point, an E coli requirement can be set for the
treated wastewater. This is best attained by waste stabilization ponds, or other
treatment processes supplemented by maturation ponds (Chapter 12).
Effluent Quality 53
Forty per cent of the world’s population lives within 60 km of the sea, and
thus the potential for adverse human impacts on the marine environment is
very high (Jasuja, 2002*). These occur because wastewaters from coastal cities
and towns are discharged, often with no treatment at all, into the sea or an
estuary via an outfall pipe. To minimize these impacts, the outfall pipe should
be correctly positioned. Its actual position in any one location depends largely
on the pattern of the local tidal currents (see Grace, 1978; Gunnerson, 1988*;
Institution of Civil Engineers, 1989; Water Research Centre, 1990, 1995;
National Research Council, 1993*; Neville-Jones and Chitty, 1996), but it
should be chosen so as to:
always discharge the wastewater below mean low water level,
ensure that there is no increased health risk to swimmers,
prevent the fouling of beaches with wastewater solids of recognizable
origin, and
minimize damage to the marine ecosystem (particularly coral reefs).
The first three of these four criteria are especially important if there is a large
tourist industry; bathing waters should contain ≤2000 E coli/100 ml,
preferably ≤100/100 ml (Council of the European Communities, 1976*). The
fourth criterion is of long-term importance, particularly if the wastewater
contains an appreciable proportion of toxic industrial wastes. If there are
commercial shellfisheries near the outfall, then the sea water in these areas
should contain ≤10 E coli/100 ml (Council of the European Communities,
The fourth criterion is especially important as coral reefs are fragile
ecosystems with very high biodiversities. Recently the lethal ‘white pox’ disease
of Caribbean elkhorn coral, Acropora palmata, has been shown to be caused
by the common faecal bacterium Serratia marescens (Patterson et al, 2002*) –
this finding, that a human faecal bacterium is a pathogen of a marine
invertebrate, is extremely disturbing and indicates the need for effective
domestic wastewater treatment. Although discharge to sea is an easy option
for wastewater disposal for coastal towns and cities, wastewater treatment
and re-use should always be considered as an alternative solution, especially
since many coastal areas in developing countries are short of water. Moreover,
we should try not to add to, but rather reduce, the pollution of the sea, which
is a global resource of considerable economic value and ecological importance.
The Aruba Protocol
In 1983 the Cartagena Convention was drawn up ‘for the protection and
development of the marine environment of the Wider Caribbean Region’ and
in 1999 the Aruba Protocol to the Convention was developed to control
marine pollution in the Caribbean from ‘land-based sources and activities’,
54 Domestic Wastewater Treatment in Developing Countries
Table 4.4 Effluent Quality Requirements for Domestic Wastewaters
Discharged into the Marine Environment of the Wider Caribbean Region
Suspended solidsa
Faecal coliformsb
Discharge into
Class I waters
Discharge into
Class II waters
≤ 30 mg/l
≤ 30 mg/l
≤ 200 per 100 ml
≤ 150 mg/l
≤ 150 mg/l
Not applicable
a Excludes algae in waste stabilization pond effluents
b Assumes discharge is into surf zone of bathing beaches
Source: Caribbean Environment Programme (2002*)
including the discharge of domestic wastewaters (Caribbean Environment
Programme, 2002*). Effluent qualities are specified for discharge into ‘Class I’
and ‘Class II’ coastal waters (Table 4.4). Class I waters are defined as those
containing coral reefs, seagrass beds or mangroves; those which are critical
breeding, nursery or forage areas for aquatic and terrestrial life; those which
provide habitat for protected marine species; and recreational waters (ie those
used for swimming and water sports). Class II waters are all other marine areas
which are less sensitive to the impacts of domestic wastewaters.
The United Nations Environment Programme has 14 Regional Seas
Programmes, one of which is the Caribbean Environment Programme. Details
of the other programmes and their discharge requirements are given by the
United Nations Environment Programme (2003*).
BATNEEC stands for the ‘best available technology not entailing excessive
cost’ and is often the technology preferred by regulators. From their point of
view this is perhaps understandable, but actually what wastewater treatment
plant operating authorities want is CATNAP – the ‘cheapest available
technology narrowly avoiding prosecution’ (the acronym CATNAP started as
a joke, but actually it is a very apposite descriptor). In practice this means there
is no point at all in having a wastewater treatment technology that produces
an effluent quality of 5 or even 20 mg BOD/l when the standard is ≤30 mg/l –
technology overkill is as bad as regulatory overkill (and both really are bad: if
too much money is spent unnecessarily on wastewater treatment to produce
an effluent quality much better than is actually needed, then there is less money
available for other basic services, such as better health care, better education,
better public transport, and so on. Regulatory and technology overkill are in
fact worse than economically wanton: they are, to put it bluntly, stupid – but,
regrettably, not so uncommon).
Regulators have to remember that most wastewaters in most developing
countries are not treated at all. If a city installs a wastewater treatment plant,
Effluent Quality 55
but for some reason (often financial, sometimes operational or managerial) the
plant produces an effluent of, say, 40 (even 60) mg BOD/l, rather that the
required ≤30 mg/l, then should it be prosecuted? If other cities in the country
(or state, or province) discharge their wastewater untreated into local rivers
and are not prosecuted for so doing, then the answer has to be No – otherwise
there would be no incentive at all to even consider wastewater treatment.
On the basis that ‘half a loaf is better than no bread’, governments and
their regulators (and indeed bilateral and multilateral donors and lending
agencies) have to realize that partial treatment is better than no treatment.
Phasing in full treatment over several years to achieve a desired river water
quality objective is more sensible than doing nothing – ‘Rome was not built in
a day.’ Of course, there will be situations where full treatment may be required
now (perhaps in re-use schemes, for example), but this is not the general case
in developing countries.
Biochemical Oxygen Demand
Removal Kinetics
The rate at which organic matter is oxidized by bacteria is a fundamental
parameter in the rational design of biological waste treatment processes. It has
been found that BOD (biochemical oxygen demand) removal often
approximates first-order kinetics; that is, the rate of BOD removal (= the rate
of oxidation of organic matter) at any time is proportional to the amount of
BOD present in the system at that time. Mathematically this type of reaction is
written as:
= – k1L
where L is the amount of BOD remaining (= organic matter still to be oxidized)
at time t; and k1 is the first-order rate constant for BOD removal, which has
the units of reciprocal time, usually day–1.
The differential coefficient dL/dt is the rate at which the organic matter is
oxidized, and the minus sign indicates a decrease in the value of L with time.
Equation 5.1 is the differential form of the first-order equation for BOD
removal; it can be integrated to:
L = L0e–k1t
where L0 is the value of L at t = 0.
L0 is the amount of BOD in the system before oxidation occurs; it is
therefore the ultimate BOD (Chapter 1). The amount of BOD removed or
‘satisfied’ (= organic matter oxidized) plus the amount of BOD remaining (=
organic matter yet to be oxidized) at any time must obviously equal the
ultimate BOD (= initial amount of organic matter):
Biochemical Oxygen Demand Removal Kinetics 57
BOD remaining (L)
oxygen utilized (y)
Figure 5.1 Generalized BOD Curves
y = L0 – L
where y is the BOD removed at time t.
Substitution of equation 5.3 into equation 5.2 yields:
y = L0 (1–e–k1t)
Generalized BOD curves (plots of equations 5.2 and 5.4) are shown in Figure
5.1, from which the relationship between y, L and L0 is readily seen.
Equation 5.2 can also be written as:
L = L010–K1t
where K1 = k1/2.3. Because of the confusion that generally arises between K1
and k1, it is best to give the base when quoting k1 values – for example 0.23
day–1 (base e) or 0.10 day –1 (base 10). In this book all first-order rate constants
are to base e.
Ratio of BOD5 to BODu
The ratio of BOD5/BODu (= y5/L0) is given by equation 5.4 as (1 – e–5k1). As
the value of k1 for raw domestic wastewater is typically 0.23 day–1 (base e) at
20°C, the ratio BOD5/BODu in raw domestic wastewaters is ~2/3.
58 Domestic Wastewater Treatment in Developing Countries
In a wholly microbiological wastewater treatment reactor (ie one in which no
non-microbiological BOD removal takes place) the difference between the
influent COD (chemical oxygen demand )and the effluent COD (∆COD) can
be used to determine the BODu removal in the reactor. This is because COD
equals BODu (which is by definition the biodegradable COD) plus the nonbiodegradable COD (which by definition cannot change in a wholly
microbiological reactor):
CODi = (BODu)i + (non-biodegradable COD)
CODe = (BODu)e + (non-biodegradable COD)
∆COD = (BODu)i – (BODu)e
Once the COD/BOD5 ratios have been determined for both the influent and
the effluent, ∆COD can be used to estimate BOD5 removal.
Continuous flow processes
Equations 5.2 and 5.4 describe the bio-oxidation of a given quantity of organic
matter to which no further addition of organic matter is made. They represent
conditions in a ‘batch’ oxidation process. However, wastewater treatment
plants operate with a continuous inflow of raw wastewater and a continuous
outflow of treated effluent. Consider, therefore, a mass balance of BOD across
a continuously operated microbiological reactor: the quantity of organic
matter entering the reactor per day must equal the quantity leaving the reactor
per day plus that removed by bio-oxidation. If Q is the flow in m3/day and Li
and Le are the influent and effluent BOD, respectively, in mg/l (= g/m3), then:
of BOD entering = L Q
( quantity
the reactor, g/day
of BOD leaving = L Q
( quantity
the reactor, g/day
The quantity of BOD removed in bacterial oxidation is given by equation 5.1
as k1L g/m3 day where L is the BOD of the reactor contents. We will assume
that the reactor is completely mixed so that the reactor contents are identical
in every respect to the reactor effluent. Under this condition the BOD of the
reactor contents is Le. If V is the working volume of the reactor in m3, then:
of BOD removed
=k LV
( byquantity
bacterial oxidation, g/day)
1 e
Biochemical Oxygen Demand Removal Kinetics 59
LiQ = LeQ + k1LVe
1 + k1(V/Q)
The ratio V/Q is the mean hydraulic retention time (θ, days), the average length
of time a typical ‘packet’ of wastewater may be expected to remain in the
reactor before being discharged in the effluent flow. Equation 5.6 can therefore
be written as:
1 + k1θ
This equation has found direct application in the design of waste stabilization
ponds and aerated lagoons (Chapters 11, 12 and 20).
The rate constant k1 is a gross measure of bacterial activity and, in common
with almost all parameters describing microbiological growth processes, its
value is strongly temperature dependent. Its variation with temperature is
usually described by an Arrhenius equation of the form:
k1(T) = k1(20) φT–20
where k1(T) and k1(20) are the values of k1 at T°C and 20°C, respectively, and φ
is an Arrhenius constant whose value is usually between 1.01 and 1.09. Typical
values are:
Waste stabilization ponds 1.05–1.09
Aerated lagoons 1.035
φ values are themselves a function of temperature, decreasing with increasing
temperature, but they are normally sensibly constant over a 10–15 degC
interval. Thus a φ value for the temperature range 5–15 degC will not be the
same as that for the range 15–30 degC. Caution must therefore be exercised in
60 Domestic Wastewater Treatment in Developing Countries
using, in developing countries, φ values derived from wastewater treatment
processes operating in temperate climates.
As a way of avoiding the problem of φ values being only valid for a fairly
narrow range of temperature, Mara (1987) used a value of φ that varies
linearly with temperature in the range 8–35°C (see equation 11.3):
φ = (1.107 – 0.002T)T–25
The flow of wastewater through a microbiological reactor can approximate
either complete mixing or plug flow. These two flow patterns represent two
extreme or ideal conditions. In practice the hydraulic regime lies between these
two extremes and is described as ‘dispersed flow’.
Complete mixing
The influent to this ideal reactor is completely and instantaneously mixed with
the reactor contents, which are, as a result of the intense mixing, uniform in
composition throughout. The effluent is identical, therefore, in every respect
to the reactor contents. The removal of BOD is described by equation 5.7.
Plug flow
The contents of this ideal reactor flow through the reactor in an orderly
fashion characterized by the complete absence of longitudinal mixing. The
concept of plug flow may be readily grasped by imagining the wastewater, on
arrival at the reactor, to be placed in watertight ‘packets’ which then travel
along the length of the reactor – as if on a conveyor belt – with no transfer of
material from one packet to another, but with complete mixing within each
packet. Since each packet receives no additional BOD and loses none to a
neighbouring packet, the removal of BOD within each packet is essentially a
batch process, so that BOD removal in a plug flow reactor follows equation
5.2. It is, however, convenient to adopt the notation used in equation 5.7 and
rewrite equation 5.2 as:
Le = Lie–k1θ
Equation 5.10 has found direct application in the design of constructed
wetlands (Chapter 17) and biofilters (Chapter 19).
Dispersed flow
It is, of course, impossible to build a plug flow reactor in which there is no
Biochemical Oxygen Demand Removal Kinetics 61
mixing between packets; in practice some degree of longitudinal mixing always
occurs. The degree of inter-packet mixing that takes place is usually expressed
in terms of a dimensionless ‘dispersion number’ (δ), defined as:
δ = D/vl
where D is the coefficient of longitudinal dispersion, m2/h; v is the mean
velocity of travel of a typical ‘particle’ in the reactor, m/h; and l is its mean
path length, m. When there is no longitudinal dispersion (ie in the case of ideal
plug flow) δ = 0 and when there is infinite dispersion (ie complete mixing) δ =
In a dispersed flow reactor (0 < δ < ∞) in which bio-oxidation occurs as a
first-order reaction, the removal of BOD is described by the equation given by
Wehner and Wilhelm (1956*):
4a exp (1/2δ)
(1 + a) exp (a/2δ) – (1 – a)2 exp (–a/2δ)
where a = Ⱦ(1 + 4k1θδ). Equation 5.12 reverts to equation 5.10 when δ = 0
and to equation 5.7 when δ = ∞.
Figure 5.2 is a graphical display of equation 5.12 (Thirumurthi, 1969),
which shows that, for any given combination of k1 and θ, maximal BOD
removal is achieved in an ideal plug flow reactor, and minimal removal in a
completely mixed reactor of the same size. Expressed in another way this
means that, for any given value of k1 and any desired degree of BOD removal,
the required retention time is a minimum in a plug flow reactor and a
maximum in a completely mixed reactor. A plug flow reactor is therefore
always smaller than a completely mixed reactor designed to achieve the same
removal of BOD.
For δ values less than 2 (and, as shown in Figure 5.2, a δ value of 2 is quite
close to infinity), the second term in the denominator of equation 5.12 is small
and can be ignored. The equation becomes:
Le/Li = [4a/(1 + a)2]exp[(1 – a)/2δ]
A least-squares analysis for the solution of equation 5.13 for k1 and δ is given
by Esen and Al-Shayji (1999*).
Dispersion numbers can be determined directly by chemical tracer studies
with either an inorganic salt (eg lithium chloride) or a fluorescent dye (eg
rhodamine WT) (Levenspiel, 1998; see also Agunwamba, 2002*). Some typical
results are shown in Figure 5.3. As an alternative to performing tracer studies,
a reasonably good estimate of δ values for facultative and maturation waste
stabilization ponds (Chapters 11 and 12) can be obtained from the pond
62 Domestic Wastewater Treatment in Developing Countries
Dimensionless product (k1θ)
Plug flow
BOD remaining (%)
Note: The numbers adjacent to each curve are the corresponding dispersion numbers.
Figure 5.2 Thirumurthi Chart for the Wehner–Wilhelm equation
geometry, using the simple equation developed by von Sperling (1999*, 2002*,
and 2003*):
δ = (L/B)–1
where L and B are the pond length and breadth, respectively, m – that is δ is
the reciprocal of the pond’s length-to-breadth ratio (Chapter 13).
BODu or BOD5?
In equations 5.7, 5.9 and 5.12 which describe BOD removal in continuous
flow reactors, the terms Li and Le may refer to either the ultimate BOD or the
5-day BOD of the influent and effluent. In practice BOD5 is most commonly
Biochemical Oxygen Demand Removal Kinetics 63
Complete mixing ( )
High dispersion (0.2)
Low dispersion (0.002)
Plug flow (0)
Note: A slug of dye is added to the influent and its concentration in the effluent Ce is determined
at various corresponding times te. The results are plotted as the dimensionless numbers Ce/C*
and te/θ, where C* is the weight of dye added divided by the reactor volume and θ is the mean
hydraulic retention time (= reactor volume/flow rate). In the completely mixed reactor the dye is
instantaneously and uniformly distributed so that at zero time Ce = C*; exponential wash out of
the dye then follows. In the plug flow reactor all the dye appears in the effluent when te = θ.
Dispersed flow reactors behave in an intermediate fashion depending on the magnitude of their
dispersion number; two examples are shown here, for δ = 0.2 and δ = 0.002.
Figure 5.3 Typical Tracer Study Results
used because it is more easily measured; k1 is thus strictly interpreted as the
first-order rate constant for BOD5 removal.
Which model to use?
The three models – complete mixing, plug flow and dispersed flow – can cause
some confusion: which is the most appropriate to use? Sometimes the answer
is obvious: in the case of rivers (Chapter 4) and long narrow constructed
wetlands (Chapter 17), the plug flow model (equation 5.10) is clearly the most
64 Domestic Wastewater Treatment in Developing Countries
appropriate, and in the case of aerated lagoons (Chapter 20) the complete mix
model (equation 5.7) is obviously appropriate. However, for waste
stabilization ponds (Chapters 10–12) the choice is less obvious. Theoretically,
of course, the dispersed flow model (equation 5.12 or 5.13) is the best, but in
practice it is only rarely used. Most commonly the complete mix equation is
used to derive a k1 value from field data; provided that this k1 value is only
used in equation 5.7, then the results will be acceptable (it is wholly wrong to
use a k1 value derived from equation 5.7 in either equation 5.10 or equations
5.12 and 5.13). The assumption that is being made is not that the pond (or
other reactor) is completely mixed, but rather that the kinetics are reasonably
well represented by the complete mix model (Marais and Shaw, 1961).
A more rational approach has been adopted by von Sperling (1999*,
2003*), who correctly advocates the use of only the dispersed flow model.
Using data from waste stabilization ponds in Brazil, he developed the following
empirical equation for kB (the first-order rate constant for E coli, rather than
BOD, removal) as a function of pond depth and retention time:
kB = 0.92D–0.88θ–0.33
where kB is the first-order rate constant for E coli removal in a dispersed flow
reactor, d–1; and D is the pond depth, m. This equation, corrected for
temperature by an equation similar to equation 5.8, and equation 5.14 are
then used in equation 5.12 or 5.13. This important approach is discussed in
more detail in Chapter 12.
Equation 5.1 assumes that all the organic components in the wastewater are
oxidized at the same rate and that the rate of oxidation remains constant with
time. However, it is unlikely that all components of a waste so heterogeneous
in nature as domestic wastewater will be oxidized at the same rate, and it has
been frequently observed in waste stabilization ponds, for example, that as the
retention time increases the rate constant decreases.
Composite exponential
If it is assumed that different fractions of the waste are oxidized at different
rates, but that each rate constant does not decrease with time, the simple
exponential term exp(–k1t) in equation 5.4 can be replaced by a composite
exponential, so that the equation becomes (Gameson and Wheatland, 1958):
y = L0 [1 – f1 exp(–kf t) – f2 exp(–kf t) – … – fn exp(–kfnt)]
Biochemical Oxygen Demand Removal Kinetics 65
where f1, …, fn are the fractions of the waste oxidized at the rates kf1, …, kfn.
For example, a domestic wastewater in England was found to be oxidized as if
it were a mixture of components of which 40 per cent were oxidized at a rate
of 0.8 day–1, 40 per cent at 0.08 day–1 and 20 per cent at 0.008 day–1:
y = L0 (1 – 0.4e–0.8t – 0.4e–0.08t – 0.2e–0.008t)
The effluent from the wastewater treatment works at which this particular
wastewater was treated, was oxidized as if it were composed of a 40 per cent
fraction oxidizable at a rate of 0.08 d–1 and a 60 per cent fraction oxidizable
at 0.008 day–1:
y = L0 (1 – 0.4e–0.08t – 0.6e–0.008t)
These typical results show that the most rapidly oxidizable fraction was totally
eliminated during treatment and, as a result, the effluent had a higher
proportion of material that could be oxidized, only more slowly. This explains
the variation of the BOD5/COD ratio during wastewater treatment: in raw
domestic wastewater it is ~0.5, but in the effluent it is lower, often ~0.2.
Retarded exponential
If, on the other hand, it is assumed that all the components in the wastewater
are oxidized at the same rate but that the rate of oxidation decreases with time,
then equation 5.1 becomes (Gameson and Wheatland, 1958):
1 + αt
where α is a coefficient of retardation, day–1; and k1 is now defined as the firstorder rate constant at zero time.
The retarded exponential equation for a completely mixed reactor is:
1 + [k1/(1 + αθ)]θ
In his research on waste stabilization ponds in northeast Brazil, Silva (1982)
obtained the data from primary facultative ponds (see Chapter 11) given in
Table 5.1. These results indicate that, as the retention time increases, the values
of k1 from equation 5.7 decrease. Using equation 5.20 Silva found an α value
66 Domestic Wastewater Treatment in Developing Countries
Table 5.1 BOD Removal Results from Primary Facultative Ponds in
Northeast Brazil
BOD loading (kg/ha d)a Retention time (d)
Raw wastewater
Effluent from:
Pond F2
Pond F3
Pond F4
Pond F5
BOD (mg/l)
Note: a see equation 11.1
Source: Silva (1982)
of 0.052 for these facultative ponds, so that k1 was given, for a mean in-pond
temperature range of 21–28°C, by:
k1 = 0.527/(1 + 0.052θ)
Estimation of ultimate BOD
In streamflow analysis (Chapter 4) use of equation 5.4 to estimate the ultimate
BOD of a wastewater effluent results in a significant underestimation of L0
and hence in a corresponding underestimation of the oxygen demand in the
river downstream of the effluent discharge point (Borsuk and Stow, 2000*).
Equation 5.1 can be rewritten as a ‘mixed’ order equation, as follows:
= – knLn
where kn is the mixed-order rate constant for BOD removal, (mg/l)1–n day–1;
and n is the order of the reaction. Integrating equation 5.22 and using equation
5.3 yields:
y = L0 – [L01–n – knt(1 – n)]1/(1–n)
Borsuk and Stow (2000*) used Bayesian statistics to establish L0 from longterm data sets of y and t (up to 180 days) for n ≠ 1. They found the
mixed-order model much better at estimating L0 than the first-order model or
indeed the second-order model (ie n = 2); n was determined for each of four
long-term data sets and found to be in the range 1.30–4.04 (but n values are
unique to each data set, so this range in no way constrains possible values of
n), and L0 was underestimated by the first-order model in these four cases by
Biochemical Oxygen Demand Removal Kinetics 67
as much as 6–45 per cent. Borsuk and Stow’s approach should be adopted for
effluent analysis prior to streamflow calculations when there is (or it is possible
to obtain) a reasonably long-term data set (say, up to at least 60 days) and
when it is important to determine in-river dissolved oxygen concentrations
with a high level of confidence.
1 The BOD5 of a wastewater has been measured as 300 mg/l. If k1 = 0.23 day–1
(base e) what is the BODu of the waste? What proportion of the BODu would
remain unoxidized after 20 days?
From equation 5.4:
[y5 = L0 ( 1 – e–k1.5 )
L0 = y5 ( 1 – e–k15 )–1
= 300 ( 1 – e–(0.23 x 5) )–1
= 400 mg/l
From equation 5.2:
= e–k120
= e–(0.23 x 20)
= 0.01
Thus 99 per cent of the waste has been oxidized in 20 days. BOD20 is therefore
often taken as an approximation for BODu.
2 Show that the ratio of the 21/2-day, 35°C BOD to the 5-day, 20°C BOD is
approximately 1. Take φ as 1.05.
From equation 5.4:
y2.5 = L0[1 – exp (– 2.5k35)]
y5 = L0[1 – exp (– 5k20)]
But from equation 5.8:
k35 = k20 (1.05)35–20
68 Domestic Wastewater Treatment in Developing Countries
= 2.08k20
Substituting k35 in the expression for y2.5:
y2.5 = L0[1 – exp(– 2.5 x 2.08k20)]
= L0[1 – exp(– 5.2k20)]
≈ y5
Thus the BOD5 of a wastewater can be obtained in 2.5 days if the incubation
temperature is 35°C rather than 20°C – see Tool (1967). It would actually be
more convenient to measure the 3-day BOD at 30ºC which can be shown in
the way described above to be essentially equal to the 5-day 20ºC BOD (in
warm climates this obviates the need for a cooled incubator).
Domestic Wastewater Treatment
The most commonly quoted definition of sustainable development is that
‘which meets the needs of the present without compromising the ability of
future generations to meet their own needs’ (World Commission on
Environment and Development, 1987 – the ‘Bruntland Report’). For water
resources, Feitelson and Chenoweth (2002*) interpret this as ‘water resources
left for future generations should be of similar quantity and quality as those
available to current generations’.
In almost all parts of the world, but especially in developing countries,
there is a huge need for water, and the water to meet these needs is becoming
scarcer and scarcer: it is predicted that over half the world’s peoples will face
water shortages during the next 30 years (Postel, 1997; United Nations
Environment Programme, 2002*; Hunt, 2003). Agriculture consumes vast
quantities of water (~70 per cent of global water abstraction), as do many
industries, and there is an enormous, currently unfulfilled, domestic demand
for water. The development and exploitation of water resources to meet these
needs must be sustainable (as defined above), and part of this drive towards
sustainability concerns domestic wastewater treatment. This includes the direct
re-use of treated wastewater in agriculture and aquaculture (Chapters 21 and
22), as well as its indirect re-use, which is discharge into inland surface waters
and used by downstream communities for agricultural and industrial use, as
well as for domestic supply.
When sustainability is considered in relation to domestic wastewater
treatment in developing countries, the following issues are relevant:
low cost – both in terms of capital and of operation and maintenance,
simplicity of operation and maintenance,
low, preferably zero, energy usage – essential for low operational costs,
low, preferably zero, use of chemicals, especially chlorine or other
environmentally damaging disinfectants,
low land take, although this is only occasionally really relevant,
70 Domestic Wastewater Treatment in Developing Countries
Table 6.1 Comparison of Factors of Importance in Wastewater Treatment in
Industrialized and Developing Countries
Sludge production
Land requirements
Environmental impact
Operational costs
Construction costs
Industrialized countries
Developing countries
Notes: C, critical; •••••, extremely important •, no impact
Source: adapted from von Sperling (1996a*)
high performance – the ability to produce an effluent of the required
quality (Chapter 4), and
low sludge production.
These considerations should be self-evident, yet they are often not taken into
account – unfortunately there are too many uninformed professionals in
developing countries (and this includes their often expatriate advisers and
lenders) who, automatically and thus without due thinking, wish to adopt the
wastewater treatment technologies of industrialized countries in the generally
mistaken belief that these are the most appropriate technologies to implement.
This belief is generally mistaken because, as shown in Table 6.1, industrialized
and developing countries have (or should have) different perceptions of what
is important in wastewater treatment. Of course, there are certain
circumstances when such technologies may be appropriate – in ‘megacities’,
for example, but even here other technologies, such as waste stabilization
ponds and effluent re-use, are not always irrelevant (see Chapter 9 for
examples of large-scale waste stabilization pond systems).
Decentralized wastewater treatment
In most cities in industrialized countries there is a single central wastewater
treatment plant. Such a plant requires an extensive (and expensive) network of
trunk sewers to convey all the city’s wastewater to it, and this often involves
pumping the wastewater from one drainage basin to another. A cheaper
alternative for developing countries (which is also applicable to industrialized
countries) is to have decentralized wastewater treatment plants, rather than a
single central plant (Lens et al, 2001*). This minimizes the costs of trunk
sewerage and avoids much, if not all, of the expenditure on pumping. Each
decentralized plant serves a single drainage basin or a small number of
drainage sub-basins. A good example of this is in Lusaka, Zambia: the city is
Domestic Wastewater Treatment Options 71
divided into six catchment areas, each of which has its own treatment plant
(Wamukwamba and Share, 2001*). Another example is the city of Curitiba
(population ~2 million) in south Brazil, where having decentralized wastewater
treatment achieved cost savings of ~40 per cent compared with having a single
treatment plant (Catunda and van Haandel, 1996*). When sewering cities for
the first time, or when an existing sewerage system has to be extended to deal
with new housing areas, consideration should always be given to decentralized
wastewater treatment and re-use as they can be more cost-effective.
In this book the following processes for the treatment of domestic wastewater
are discussed:
waste stabilization ponds (WSP) (Chapters 9–15),
wastewater storage and treatment reservoirs (WSTR) (Chapter 16),
constructed wetlands (CW), often simply called ‘reedbeds’ (Chapter 17),
upflow anaerobic sludge blanket reactors (UASBs) (Chapter 18),
biofilters (Chapter 19),
aerated lagoons (Chapter 20), and
oxidation ditches (also Chapter 20).
Not all of these technologies are necessarily sustainable or always sensible.
Some may not be sensible at all, but are included here because they are often
advocated as being a good (sometimes even ‘the best’) solution. Professionals
(and would-be professionals) need to understand these technologies so that
they are able to make informed decisions about which is really the best
technology, or combination of technologies, to implement in any given
situation, and also which technologies to avoid.
Conventional activated sludge systems are not considered in this book,
although two of its variants are – aerated lagoons and oxidation ditches
(Chapter 20). Conventional systems are described in, for example, Horan
(1990), Metcalf and Eddy, Inc (1991) and Water Environment Federation
(2001). They should only be used for very large populations (‘megacities’),
and then only with extreme caution as they consume considerable quantities
of electrical energy, are very complicated to operate and maintain (so highly
skilled operators are needed), and generally require very large amounts of
foreign exchange, both to equip initially and then to maintain. The Latin tag
caveat emptor (‘let the buyer beware’) is highly relevant as sales engineers
selling conventional wastewater treatment equipment are highly skilled in
Lime-assisted primary sedimentation is a good method of chemically
enhanced primary treatment of domestic wastewater, although quite high doses
of lime are required (up to 1 g/l) and large amounts of sludge are produced
(around 0.14 m3/m3 wastewater treated) (Gambrill, 1990; Taylor et al,
72 Domestic Wastewater Treatment in Developing Countries
1994a*, b*; see also Jiménez-Cisneros et al, 2001*; Harleman and Murcott,
1999*; Environmental Protection Agency, 2000a*). At retention times of 9–12
h and pH 11, Gambrill (1990) found that helminth eggs were removed by 4
log units and faecal coliforms by 4.5 log units, so producing an effluent safe
for unrestricted irrigation (Chapter 21). In regions with acid soils the lime
sludge can be profitably used to reduce soil acidity (and, although the lime
sludge will contain many Ascaris eggs, almost all will be damaged by the lime
and consequently will be inviable, so not posing a health risk). Lime sludge
disposal to even alkaline soils does not raise the soil pH significantly, and yields
of cotton, for example, from lime sludge-amended plots were found to be
higher than those from control plots (Akrivos et al, 2000*; see also Jiminez et
al, 2002).
Membrane bioreactors are a relatively recent wastewater treatment
technology. They can produce extremely high quality effluents (see Ben Aim
and Semmens, 2003*) – but usually of too high a quality for, and therefore
generally inappropriate in, most situations in almost all developing countries
at present. Details are given in Stephenson et al (2000) and van der Roest et al
(2002); see also Chang et al (2002*) on membrane fouling in these systems.
Many developing countries have a warm or hot climate, and often they have
sufficient land for land-intensive wastewater treatment technologies (WSP, for
example). They should take maximal advantage of their climate and their land
availablilty for wastewater treatment. Money spent on land is not money
wasted (Chapter 9), but money spent on electricity is money gone for ever.
Thus the most sustainable options for domestic wastewater treatment in
developing countries are likely to be:
anaerobic technologies, such as anaerobic ponds (especially high-rate
anaerobic ponds) and maybe UASBs, and
photosynthetic technologies, such as facultative and maturation ponds and
maybe constructed wetlands.
Generally these two ‘natural’ processes are used in series, treating the
wastewater first anaerobically and then photosynthetically (see also Gijzen,
2002*). Given modern design the land take for these natural systems need not
be as large as their ‘opponents’ commonly suppose or might suggest. One
example will suffice here (the point is considered in more detail in Chapter 9):
WSP have been described as suitable ‘only if land is relatively cheap (<US$
15/m2)’ (Yu et al, 1997). This translates to WSP being suitable at land prices
up to US$150,000/ha, which is in fact a very high price for land near almost
all towns and cities in developing countries – the best agricultural land in
industrialized countries is worth much less than this (in England, for example,
the very best ‘general purpose arable’ land costs up to ‘only’ US$20,000/ha).
Domestic Wastewater Treatment Options 73
If a move is made away from these natural treatment processes (or forced
away, due to inappropriately high effluent quality requirements – Chapter 4),
then there is in effect a trade-off: more money needs to be spent on both capital
and operation and maintenance (O&M) costs (especially electrical energy
costs), and less on land. In simple financial terms, such as a discounted cashflow analysis (ie in average incremental cost terms), it is basically almost a
straight comparison between land costs and the costs of electromechanical
equipment (including their concrete housing structures) and the electricity
used. As will be shown in subsequent chapters (especially Chapter 9), natural
systems are almost always the most appropriate option. Higher-tech systems
are only appropriate when the land for natural systems really is not there. Then
one should move up the higher-tech ‘ladder’ very cautiously, always
considering the least higher-tech options first and the most high-tech systems
last. It will often be instructive, especially for large wastewater treatment
projects, and particularly when high-tech solutions are being considered, to
undertake a Bayesian benefit–risk analysis to determine the most sustainable
option (Englehardt, 1997*).
Domestic Wastewater Flows and
Domestic wastewater flows are commonly determined from domestic water
Qww = 10–3kqP
where Qww is the wastewater flow, m3/day; q is the water consumption,
l/person day; P is the population connected to the sewerage system; and k is
the ‘return factor’, the fraction of the water consumed that becomes
wastewater. The value of k is usually 0.8–0.9. It is lower in rich areas where
water is used for car washing and garden watering. Equation 7.1 gives the
domestic wastewater ‘dry weather flow’ (DWF) – a term used principally from
the time when ‘combined’ sewers (ie sewers receiving both sanitary and
stormwater flows) were common. (Combined sewers do exist in developing
countries, especially in city centres, but the current preference is to separate
sanitary and stormwater flows.) Dry weather flow is the average wastewater
flow per day over seven consecutive days without any rain which follow seven
days with no more than 0.25 mm of rain on any one day. The mean daily flow
is often taken as 1.3 x DWF.
Butler and Graham (1995*) describe a computer model called ‘Flush’ to
determine actual DWF values (and also peak flow values – see below) based
on detailed surveys of in-house water usage, and also to assess the spatial and
temporal variations in DWF.
Campos and von Sperling (1996*) give the following regression equation
for the mean domestic wastewater flow (qww, l/person day) based on household
income in Brazil:
qww = 58 + 8Nms
Domestic Wastewater Flows and Loads 75
where Nms is the household income expressed as the number of minimum
salaries per month (one minimum salary was US$100/month in Brazil in 1996).
Equation 7.2 is valid only for Brazil and perhaps other Latin American
countries that are socio-economically broadly similar to Brazil. The derivation
of similar equations for African and Asian developing countries would be very
Sewer joints are often imperfect; over time they allow groundwater to enter
the sewer. This increases the wastewater flow (and decreases the wastewater
BOD). For concrete pipes, infiltration is ~20 m3/ha day; and for PVC pipes
(which are easier to joint well) ~10 m3/ha day. Thus taking infiltration into
account, equation 7.1 becomes:
Qww = 10–3kqP + I
where I is the infiltration flow, m3/day.
Industrial wastewaters
In a municipal wastewater treatment project the flow of industrial wastewater
must also be taken into account. Equation 7.3 becomes:
Qww = 10–3kqP + I + E
where E is the industrial wastewater flow, m3/day.
Many industrial wastewaters are toxic to microbiological wastewater
treatment processes, and therefore they should be pretreated prior to discharge
to sewer (see Water Environment Federation, 1994a). Failure to pretreat can
be extremely serious, and regulators need to work closely with industries to
ensure, first by persuasion and later by the imposition of ‘pollution charges’,
that they do pretreat their wastewaters (see Afsah et al, 1996*; D’Arcy et al,
1999*; Wheeler, 2000*; Tilche and Orhon, 2002*; World Bank, 2002*).
In many developing countries, industrial wastewaters are discharged
untreated into inland or coastal waters – with huge adverse impacts on the local
aquatic ecology. The 2002 Stockholm Statement on global water security
strongly recommends that ‘the link between economic growth and water
degradation be urgently broken’ (Stockholm International Water Institute,
2002*). To break this link governments and their regulators must require, either
through new legislation or the enforcement of existing legislation, that
industrial wastewaters are pretreated before discharge to sewer or properly
treated before discharge to inland or coastal waters – and this also applies, of
course, to domestic wastewaters since economic growth generally leads to
increased urban populations and hence to increased domestic wastewater flows.
76 Domestic Wastewater Treatment in Developing Countries
COD (mg/l)
Flow (m3/d)
Time (2/3 January 1973)
Figure 7.1 Diurnal Variation of Wastewater Flow (ɂ) and Load (ɀ) at
Nakuru, Kenya
Peak wastewater flows
Equations 7.1–7.4 give the mean daily wastewater flow. However, wastewater
flows (and loads) vary throughout the day (Figure 7.1), also through the week
and the year. Flows are low at night when people are asleep; they rise sharply
around breakfast time; there may be a similar ‘peak’ at lunchtime and then
one again in the evening. Certain household activities may be more common
on weekdays, rather than at weekends (eg clothes washing). Finally, during the
hot season, people may use more water (eg through showering more
It is necessary to determine the peak daily flow as this is the maximum
flow that arrives at the treatment works and parts of the works must be
designed to receive this flow (eg the preliminary treatment units detailed in
Chapter 8). The peak daily wastewater flow is the mean daily flow multiplied
by a ‘peak factor’; this depends on the population served, since the higher the
population, the lower the peak factor as flow fluctuations are reduced during
the time of flow in the sewer. The peak factor (PF) can be estimated from:
PF = 14P–1/6
where P is the population served. Equation 7.5 gives values of ~3 for a
population of 10,000, ~2 for 100,000 and ~1.4 for 1,000,000.
Domestic Wastewater Flows and Loads 77
The BOD concentration of domestic wastewater is the BOD contribution in
mg/person/day divided by the wastewater flow in l/person day. As noted in
Chapter 1, the BOD contribution increases with income. Campos and von
Sperling (1996*) give the following regression equation for BOD concentration
(Li, mg/l) based on household income in Brazil:
Li = 247 + exp(5.9 – 0.26Nms)
As with equation 7.2, this equation is valid only for Brazil and perhaps other
Latin American countries. Again similar equations for African and Asian
developing countries would be very useful.
Urban populations will certainly increase, and thus wastewater flows and loads
will also increase with time. The following equation is used to predict
population growth:
Pn = P0(1 + r)n
where Pn and P0 are the sewered populations in n years time and now,
respectively; and r is the annual population growth rate expressed as a decimal
fraction (rather than as a percentage).
The value of P0 and the anticipated value of r should be obtained from the
local municipal planning department. They should be consistent with values
from the recent past, but they also need to take into account any major
developments expected to occur in the planning period of n years. Often n is
taken as 20 years, but it is best to determine P in steps of 5 years (ie for n = 5,
10, 15 and 20 years) in order to decide how best to phase the development of
the proposed wastewater treatment facilities. Phasing is important since, rather
than building a waste stabilization pond complex to serve the population
expected in 20 years time, it is financially more sensible to build it for the
population expected in, say, 5 years time and then expand it in 4 years time to
serve the population anticipated in 10 years time, and so on. However, there
are two things that must be done now: one is to buy all the land required to
serve the population expected in 20 years time; and the other is to arrange
with the lending agency that the total loan is to be disbursed over time – some
now and some more in 4, 9 and 14 years time.
As sewered populations become richer, their per caput water consumption,
and hence their per caput wastewater generation, and their per caput BOD
contribution will increase (as shown by equations 7.2 and 7.6). This must also
be taken into account.
Preliminary Treatment
The first stage of wastewater treatment is the removal of large floating objects
(such as rags, maize cobs, pieces of wood) and heavy mineral particles (sand
and grit). This is done in order to prevent, for example, floating material
accumulating on the surface of waste stabilization ponds and heavy solids
entering the pond sludge layer, and to protect from damage the equipment
used in the subsequent stages of treatment (for example the floating aerators
in aerated lagoons or any pumps which may be used). This preliminary
treatment comprises screening and grit removal.
Manuals of Practice on preliminary wastewater treatment have been
prepared by the Institution of Water and Environmental Management (1992)
and the Water Environment Federation (1994b).
Coarse solids are removed by a series of closely spaced mild steel bars placed
across the flow. The velocity through the screen should be <1 m/s so that the
solids already trapped on the screen (the ‘screenings’) are not dislodged. The
spacing between the bars is usually 15–25 mm and the bars are commonly of
rectangular cross-section, typically 10 x 50 mm. At small works screens are
raked by hand and, in order to facilitate this, the screens are inclined,
commonly at 60° to the horizontal (Figure 8.1). The submerged area of the
hand-raked screens is calculated on the empirical basis of 0.15–0.20 m2 per
1000 population; this assumes that the screens are raked at least twice each
For flows >1000 m3/day mechanically raked screens (Figure 8.2) are
preferred since they can be cleaned more frequently (every 10–30 minutes) and
they are therefore considerably smaller than hand-raked screens. The channel
dimensions required for a mechanically raked screen are calculated as follows:
flow area =
Preliminary Treatment 79
Figure 8.1 Simple Manually Raked Screen (flow is from left to right)
The flow area is the channel area corrected for the area of the bars. The flow
is the daily maximum (ie ‘peak’) flow (Chapter 7). The velocity is generally
restricted to 0.6 m/s in order to prevent grit deposition and dislodgement of
screenings. The equation is therefore:
WDmax[s/(b + s)] = Qmax/0.6
that is
W = Qmax/{0.6Dmax[s/(b + s)]}
where W and Dmax are the channel width and depth at maximum flow,
respectively, m; Qmax is the maximum (ie peak) flow, m3/s; b and s are the bar
thickness and spacing, respectively, in millimetres.
A standby hand-raked screen should be provided for use when the
mechanical screen is out of action. This emergency screen is normally the same
size as the mechanical screen and it therefore requires raking at frequent
intervals when in use.
Disposal of screenings
Screenings are generally obnoxious in both appearance and content and should
be disposed of as soon as possible. At small works this is readily achieved by
burial, a small area of land being set aside for this purpose. At larger works
80 Domestic Wastewater Treatment in Developing Countries
Figure 8.2 Mechanically Raked Screen
screenings are commonly dewatered in a hydraulic press and then either buried
(or sometimes incinerated) on-site or taken away to the nearest landfill. Advice
on handling screenings in given by Clay et al (1996*).
The quantity of screenings that is removed varies considerably but, for 10
mm bars at 20 mm spacings, an approximate figure is 0.01–0.03 m3/day per
1000 population.
Fine screening
Fine screens have apertures of 3–15 mm, with 6 mm being the most common,
and very fine screens (‘milli-screens’) are those with apertures of 0.25–3 mm.
They produce very large quantities of screenings which are normally washed
and dewatered before disposal; these processes are generally an integral part
of the fine screen unit – for example, Filtech or Wash-flow screens (Jones and
Attwood, 2002*).
Fine screens are now common in many industrialized countries, but in
developing countries their applicability is much more limited: their high
efficiency is generally too high (‘technology overkill’ – Chapter 4), they are
expensive imported items, and their maintenance is likely to be problematic.
Preliminary Treatment 81
Grit (also called ‘detritus’) is the heavy inorganic fraction of the wastewater
solids. It includes road grit, sand, eggshells, ashes, charcoal, glass and pieces of
metal; it may also contain some heavy organic matter such as seeds and coffee
grounds. Grit has an average relative density of ~2.5 and thus it has a much
higher settling velocity than organic solids (~30 mm/s, compared with
~3 mm/s). This difference in sedimentation rates is exploited in grit removal
plants where, for ease of handling and disposal, the organic fraction must be
kept to a minimum (<15 per cent). There are two basic types of grit removal
plant: constant velocity grit channels and the various proprietary grit tanks or
traps available commercially.
Constant velocity grit channels
If the velocity of flow of the wastewater is ~0.3 m/s, grit particles settle out
but organic solids do not. The problem is to maintain the velocity constant at
this value for all rates of flow. The best solution is to locate a standing wave
(ie Venturi or Parshall) flume immediately downstream of a grit channel of
parabolic cross-section (Townsend, 1937; Marais and van Haandel, 1996*).
This solution depends on the following two points:
provided that it is free-flowing (ie not ‘drowned’), a Venturi or Parshall
flume produces an upstream depth which is independent of conditions
downstream and controlled only by the magnitude of the flow; and
if the geometry of the grit channel is such that its cross-sectional area is
proportional to the flow, then the velocity of flow through the channel will
be constant at all flows – if v = velocity, q = flow and a = cross-sectional
area, then v = q/a; but if a is proportional to q (ie a = kq) , then v is
In order to comply with (2) the channel should have a parabolic cross-section.
The explanation for this is as follows:
(a) the flow q through a Venturi flume is given by:
q = kbh3/2
where k is a constant; b is the flume throat width; and h is the upstream
channel depth.
(b) differentiating equation 8.3:
dq = 3 kbh /2dh
82 Domestic Wastewater Treatment in Developing Countries
(c) the flow dq through a cross-sectional element of the channel (Figure 8.3) is
given by:
dq = Vxdh
where V is the velocity of flow and xdh is the area of the element.
(d) equating equations 8.4 and 8.5 and rearranging gives:
x =
( 3kb
2V )
Equation 8.6 is the equation of a parabola. In practice, for ease of
construction, a trapezoidal cross-section is used (Figure 8.4).
If V = 0.3 m/s and X and H are the channel dimensions (m) at maximum
flow (Qmax, m3/s), equations 8.3 and 8.6 can be rewritten as:
Qmax = kbH
X = 5kbH
Figure 8.3 Flow Elements in a Parabolic Channel
Preliminary Treatment 83
300 mm min.
Figure 8.4 Trapezoidal Approximation to a Parabolic Section
Dividing equation 8.8 by equation 8.7 and rearranging gives:
X = 5Qmax/H
Thus the top width of the channel is simply determined from the maximum
flow and the corresponding depth. In practice at least two channels are
provided so that one may be closed for grit removal. The channel length is
determined by the settling velocity of the grit particles.
Length of channel =
channel depth x velocity of flow
settling velocity of grit particles
Grit particles typically settle at about 0.03 m/s, so that when the velocity of
flow is controlled to 0.3 m/s thus:
Length of channel = 10 x maximum depth of flow
In practice, to allow for inlet turbulence and variations in settling velocity, the
channel length is taken as 20 x maximum depth of flow.
84 Domestic Wastewater Treatment in Developing Countries
Proprietary grit separators
For flows >5000 m3/day proprietary grit separators are often more economical
than several long constant-velocity grit channels. There are several models
available. One of the simplest is the ‘Jeta’ grit trap which has the advantages
that no moving parts come into contact with the grit and that the grit is
automatically cleaned before discharge (Jones and Attwood, 2002*).
Grit disposal
The quantity of grit collected is usually in the range 0.05–0.10 m3/1000 m3 of
wastewater. The grit is either buried on site or taken to landfill.
The wastewater flow should always be measured in a Venturi or Parshall flume
before it enters the treatment reactors (if constant velocity grit channels are
used, such a flume is required anyway – see above; if not, then one must be
installed). Flow measurements are very useful for determining diurnal flow
variations (Chapter 7) and detecting any abnormal flow rates; they are also
essential for evaluating the performance of the treatment system (Chapter 15).
Digital flow recorders are useful as the data can be downloaded to a computer.
Waste Stabilization Ponds
Waste stabilization ponds (WSP) (Figure 9.1) are large shallow basins enclosed
by earth embankments in which raw wastewater is treated by entirely natural
processes involving both algae and bacteria. Since these processes are unaided
by wastewater treatment engineers (who merely allocate a properly
dimensioned place for their occurrence) the rate of oxidation is slower, and as
a result hydraulic retention times are longer than in conventional wastewater
treatment (eg electromechanical treatment processes such as activated sludge)
_ retention times in WSP are measured in days rather than in hours. They are
without doubt the most important method of wastewater treatment in
developing countries where sufficient land is normally available and where the
temperature is most favourable for their operation. Indeed they are so
advantageous (see below) that a very good case has to be made for not using
There are three principal types of WSP: anaerobic, facultative and
maturation ponds. Anaerobic ponds (Chapter 10) and facultative ponds
(Chapter 11) are designed for BOD (biochemical oxygen demand) removal,
and maturation ponds (Chapter 12) are designed for faecal bacterial removal.
Some removal of faecal bacteria (especially of Vibrio cholerae) occurs in
anaerobic and facultative ponds, which are also responsible for most of the
removal of helminth eggs; and some removal of BOD occurs in maturation
Figure 9.1 One of Phase II 21-ha Primary Facultative Ponds at Dandora,
Nairobi, Kenya
86 Domestic Wastewater Treatment in Developing Countries
ponds, which also remove some of the nutrients (N and P). There are two other
types of WSP: macrophyte ponds and ‘advanced pond systems’ which are
described later in this chapter (but they are not generally applicable in
developing countries).
Facultative and maturation ponds are photosynthetic ponds – that is the
oxygen needed by the pond bacteria to oxidize the wastewater BOD is mainly
supplied by micro-algae that grow naturally and profusely in these ponds (and
thus give them their characteristic green colour); and the carbon dioxide
needed by the algae is mainly provided by the pond bacteria as an end-product
of their metabolism. Thus there is a ‘mutualistic’ relationship between the pond
algae and the pond bacteria (Figure 9.2). The algae are also extremely
important in creating conditions within the ponds for faecal bacterial die-off
(Chapter 12).
Arrangement of WSP
The different WSP types are arranged in a series – first an anaerobic pond,
then a facultative pond, and finally (and if needed to achieve the required
effluent quality – Chapter 4) one or more maturation ponds (Figure 9.3). At
any one site there may be more than one series of WSP, each usually receiving
an equal proportion of the wastewater flow.
It is commonly observed that the effluent from a series of ponds is of better
quality than that from a single pond of the same size. This is because, even if
the hydraulic flow regime in individual ponds is closer to complete mixing
than it is to plug flow, the overall performance of a series of ponds
New cells
New cells
Figure 9.2 Algal–bacterial Mutualism in Facultative and Maturation Ponds
Waste Stabilization Ponds 87
Figure 9.3 Typical WSP Layout: A, anaerobic pond; F, facultative pond;
M1–Mn, maturation ponds
approximates that of a plug flow reactor. To illustrate this point, consider a
series of n identical ponds and assume that each is a completely mixed reactor
in which BOD removal follows first-order kinetics (Chapter 5). For the whole
series of n ponds, equation 5.7 is rewritten as:
(1 + k1θ)n
where Le and Li are the BOD of the final effluent and raw wastewater,
respectively, mg/l.
If σ is the total retention time in the series of n ponds, then σ = nθ and:
(1 + k1θ)σ/θ
In the limit as θ 0 (and if σ is to remain the same as n ∞):
(1 + k1θ)σ/θ
that is:
= exp (–k1σ)
Equation 9.3 is the equation for a plug flow reactor of retention time σ
(compare equation 5.10). This shows that a plug flow reactor is equivalent to
a series of an infinitely large number of infinitely small completely-mixed
reactors. Since plug flow is the most efficient hydraulic regime (Chapter 5),
equation 9.3 demonstrates that a series of small ponds is more efficient than a
single large pond.
88 Domestic Wastewater Treatment in Developing Countries
Marais’ theorem
This theorem, propounded by Marais (1974), states that ‘maximum efficiency
in a series of ponds is achieved when the retention time in each pond is the
same’. The theorem is proved here for a series of two ponds, but the proof is
applicable in principle for any number of ponds.
Let the retention time in the first pond be θ1 and in the second pond θ2,
and let σ = θ1 + θ2. The effluent BOD from the second pond is given by:
Le =
(1 + k1θ1) (1 + k1θ2)
Le is a minimum when [(1 + k1θ1)(1 + k1θ2)] is a maximum:
Let y = (1 + k1θ1)(1 + k1θ2)
∴ y = [1 + k1θ1] [1 + k1(σ – θ1)]
= 1 + k1σ + k12σθ1 – (k1θ1)2
= k21σ – 2k21θ1
= 0, for a maximum
∴ θ1 = σ/2 = θ2
Thus, for maximum y and hence minimum Le, the retention time in the two
ponds must be the same. That y is a maximum (and not a minimum) is verified
by considering the second differential coefficient:
= – 2k21
Since this is negative, y is a maximum.
However, in practice it is not possible for all ponds in a series to have the
same retention time. Anaerobic ponds are sized on the basis of volumetric
BOD loading (Chapter 10) and facultative ponds on the basis of surface BOD
loading (Chapter 11), and these design procedures result in different retention
times which, if they were altered to be the same, would result in an
underloaded anaerobic pond and an overloaded facultative pond – clearly
Waste Stabilization Ponds 89
undesirable. But maturation ponds (Chapter 12) should be equally sized
wherever (as is usually the case) the site geometry permits this.
The advantages of WSP are that they are simple, low-cost, highly efficient and
WSP are simple to construct: earthmoving is the principal activity; other civil
works are essentially minimal – preliminary treatment, inlets and outlets, pond
embankment construction and protection and, if necessary, pond lining (details
are given in Chapter 13). They are also very simple to operate and maintain:
routine tasks comprise cutting the embankment grass, removing any scum and
floating vegetation from the pond surfaces, keeping the inlets and outlets clear,
and repairing any damage to the embankments (details are given in Chapter
14). Less skilled labour is needed for pond operation and maintenance (O&M)
than other wastewater treatment technologies. The simplicity of WSP
construction also means that flexibility in construction phasing is possible
(Chapter 7).
Low cost
If an honest cost comparison is made between WSP and other wastewater
treatment options, WSP almost always cost the least, for both capital and
O&M costs. The only sensible methodology for such a cost comparison is that
described in a World Bank report by Arthur (1983*), which gives a detailed
economic comparison of WSP, aerated lagoons, oxidation ditches and
biofilters. The data for this cost comparison were taken from the city of Sana’a
in the Yemen Arab Republic, but are equally applicable in principle to other
countries. Certain assumptions were made: for example, the use of maturation
ponds to follow the aerated lagoon, and chlorination of the oxidation ditch
and biological filter effluents, in order that the four processes would have an
effluent of similar bacteriological quality so that fish farming and effluent reuse for irrigation were feasible. The design was based on a population of
250,000; wastewater flow and BOD contributions of 120 l and 40 g/person
day, respectively; influent and required effluent E coli numbers of 2 x 107 and
≤1 x 104 per 100 ml, respectively (but we would now design for ≤1000 E coli
per 100 ml of effluent); and a required effluent BOD of ≤25 mg/l. The
calculated land area requirements and total net present cost of each system
(assuming an opportunity cost of capital of 12 per cent and a land cost of
US$5/m2) are shown in Table 9.1. WSP were clearly the cheapest option. The
preferred solution is, of course, very sensitive to the price of land, and the
above cost of US$5/m2 was chosen as it represented a reasonable value for
land in low-cost housing estates in developing countries.
90 Domestic Wastewater Treatment in Developing Countries
Table 9.1 Costs and Land Area Requirements for WSP and other
Treatment Processes
Costs (million US$)
Benefits (million US$)
Irrigation income
Pisciculture income
Net present cost
(million US$)
Land area (ha)
pond system
Source: Arthur (1983*)
The price of land and the opportunity cost of capital (OCC) are the two main
variables which decide whether WSP are the cheapest option. Figure 9.4, based
on Arthur’s (1983*) data, shows how the OCC (or discount rate) influences
the land price below which WSP are the cheapest option. In Arthur’s study this
was US$5–15/m2 (ie US$50,000–150,000/ha), depending on the OCC within
a range of 5–15 per cent. Tsagarakis et al (2003*) found that WSP were the
cheapest treatment option in Greece up to a land cost of US$300,000/ha.
WSP do not need electrical energy for their operation, and this saves a
considerable amount of recurrent costs. This is well illustrated by the following
data for energy consumption by four wastewater treatment processes in the
USA for a domestic wastewater flow of 1 million US gallons/day (ie 3780 m3/d)
(Middlebrooks et al, 1982):
Activated sludge
Aerated lagoons
1 000 000 kWh/year
800 000 kWh/year
120 000 kWh/year
As noted in Chapter 6, money spent on electricity is money gone for ever, but
money spent on land is an investment, for example, the city of Concord,
California purchased land for ponds in 1955 at a cost of US$50,000/ha, and
by 1975 it was worth US$375,000/ha (Oswald, 1976). Inflation during this
20-year period was more or less 100 per cent, so the land value increased in
real terms by 375 per cent (or 6.8 per cent/year).
A good example of WSP investment costs is from Colombia, where the
Ministry of Economic Development uses the following budget costs:
Waste Stabilization Ponds 91
Land price (US$ per m2) below which waste stabilization
ponds are cheapest tratement option
Discount rate (%)
Source: Arthur (1983*)
Figure 9.4 Variation of Discount Rate (ie opportunity cost of capital) and
Land Price below which WSP are the Cheapest Treatment Option
Activated sludge
Colombian cities of <250,000 population are required by the Ministry to
undertake a ‘serious’ evaluation of WSP as a wastewater treatment option
(otherwise they are not considered for a government loan).
High efficiency
WSP are extremely efficient: they can be easily designed to achieve BOD and
suspended solids removals >90 per cent, and ammonia removals also >90 per
cent (Table 12.1). They are particularly efficient in removing excreted
pathogens, whereas, in contrast, all other treatment processes are very
inefficient in this respect and require a tertiary treatment process such as
chlorination (with all its inherent operational and environmental problems) to
achieve the destruction of faecal bacteria. Activated sludge plants may, if
operating very well, achieve a 99 per cent removal of E coli – this might, at
first inspection, appear very impressive, but in fact it only represents a
reduction from ~107 per 100 ml to ~105 per 100 ml (ie in terms of bacterial
92 Domestic Wastewater Treatment in Developing Countries
Table 9.2 Excreted Pathogen Removals in WSP and Conventional
Treatment Processes
Excreted Pathogen
Removal in WSP
Removal in conventional treatment
Protozoan cysts
Helminth eggs
up to 6 log units
up to 4 log units
1–2 log units
1–2 log units
Note: 1 log unit = 90 per cent removal; 2 = 99 per cent; 3 = 99.9 per cent, and so on
numbers, almost nothing). A properly designed series of WSP, on the other
hand, can easily reduce E coli numbers from ~107 per 100 ml to <103 per 100
ml (the WHO guideline for unrestricted irrigation; see Chapters 4 and 21),
which is a removal of 99.99 per cent (or 4 log units). A general comparison
between WSP and conventional treatment processes for the removal of
excreted pathogens is shown in Table 9.2; detailed information is given in
Feachem et al (1983*).
WSP are very robust: due to their long hydraulic retention time, they are more
resilient to both organic and hydraulic shock loads than other wastewater
treatment processes. They can also cope with high levels of heavy metals, up
to 60 mg/l (Moshe et al, 1972; see also Toumi et al, 2000*), so they can treat
a wide variety of industrial wastewaters that would be too toxic for other
treatment processes. Strong wastewaters from agro-industrial processes (eg
abattoirs, food canneries, dairies) are easily treated in WSP. Moreover, WSP
are the only secondary/tertiary treatment process which can readily and
reliably produce effluents that are microbiologically safe for re-use in
agriculture and aquaculture (Chapters 21 and 22).
The National River Conservation Directorate (2003*), which is part of the
Ministry of Environment and Forests, Government of India, clearly recognizes
the advantages of WSP, as it states that ‘only waste stabilization ponds, which
are eco-friendly and simple to operate, will be mainly supported to treat
wastewater hereafter’. The principal reason for this major shift in policy away
from capital- and energy-intensive treatment technologies was the realization
that WSP are not only low-cost but are also the most appropriate treatment
system to reduce pathogens to levels suitable for safe re-use in agriculture and
aquaculture and for safe bathing in the receiving rivers. NRCD considers WSP
to be ‘people-friendly and relevant for health’.
Waste Stabilization Ponds 93
To the unthinking, WSP seem to have the ‘disadvantage’ of simplicity, which is
interpreted wholly erroneously as a lack of sophistication – ‘activated sludge
must be better’! Engineers at least should know better, but unfortunately there
are too many who do not. This lack of intellectual critical ability sometimes
appears in those who really should be more aware, for example, it was
suggested at the Fifth IWA Pond Conference held in 2002 that in one southern
Pacific country, because the WSP designed according to the 1974 national WSP
design guidelines were not working well, ‘advanced pond systems’ (see later in
this chapter) should be used instead. The logical (and certainly cheaper) option
would be to upgrade the existing WSP using the modern design methods given
in Chapters 10–12.
Of course, cities and towns can be ‘turned off’ WSP by a bad experience
with them – sometimes resulting from poor operation and maintenance, or
allowing them to become seriously overloaded; but this could also be due to
poor design by an inexperienced designer or, not uncommonly, a designer
experienced with WSP in Europe or North America but who has little idea
how to design WSP in warmer climates (one of the purposes of this book is to
minimize such occurrences).
There are, however, more serious perceived disadvantages of WSP: a fear
of odour release, high land requirements and costs, and ‘poor’ effluent quality.
Odour release
All overloaded microbiological wastewater treatment plants have odour
problems, and WSP are no exception to this. However, WSP that are properly
designed and then properly operated and maintained do not have odour
problems, and they will not have odour problems until they become
overloaded. A WSP system (and this refers especially to anaerobic and primary
facultative ponds – Chapters 10 and 11) designed appropriately for a
population of N will probably be satisfactory for a population of ~1.2N (as
there is a factor of safety in the design), but it would be unrealistic not to
expect at least some odour from the ponds when they are more heavily
overloaded than this. In such cases the WSP should be upgraded (Chapter 13).
WSP do require much more land than conventional treatment processes such
as oxidation ditches (Chapter 20) or activated sludge. The choice is between
buying land (which is an investment – see earlier in this chapter) or spending a
large amount of money each year on electricity. It is quite straightforward to
determine where inexpensive land is available and how much it costs to
(a) convey the wastewater there and (b) construct WSP – is this cheaper than
an activated sludge plant nearer the city? If these cost calculations are done
correctly and honestly, the answer to this question is almost always Yes.
94 Domestic Wastewater Treatment in Developing Countries
Strange decisions are, of course, made. For example, the capital city of a
large oil-producing country between the Middle East and South Asia chose,
after a 40-year deliberation period, activated sludge over WSP – despite the
fact that it is surrounded by desert where land is cheap and the treated
wastewater is needed for crop irrigation. This example is not, unfortunately,
There will, of course, be some situations where the land for WSP is really
not available, or is prohibitively expensive, even at some distance from the
town or city. In such cases WSP are obviously inappropriate, and an alternative
treatment process has to be selected. The work of Arthur (1983*) suggests that
oxidation ditches (Chapter 20) are likely to be the next preferred solution.
Effluent quality
WSP effluents can be high in BOD and suspended solids (SS), but most of this
effluent BOD and SS is due to the algae in the pond effluent. In the BOD test
algae consume oxygen as the BOD bottles are incubated in the dark, but in a
receiving watercourse the algae produce more oxygen than they need for
respiration and they are consumed by higher life forms. Algal BOD and SS are
thus very different from non-algal BOD and SS, and sensible regulators
recognize this, for example, in the European Union the required WSP effluent
quality is ≤25 mg filtered (ie non-algal) BOD per litre and ≤150 mg SS (ie
mainly algal SS) per litre (Council of the European Communities, 1991a*). If
this is acceptable in Europe, why not in developing countries? Local regulators
in developing countries should at least be aware of these EU requirements for
WSP effluents.
If WSP effluents are used for crop irrigation (Chapter 21), the algae are
very useful: they act as slow-release fertilizers and over time increase the
organic content of the soil and thus its water-holding capacity.
WSP are used in all parts of the world – from Alaska to New Zealand. The US
has more than 7000 WSP systems (one-third of all wastewater treatment
plants), mostly serving populations of less than 5000 (Environmental
Protection Agency, 1983). France has over 2500 for populations of usually less
than 1000 (Cemagref and Agences de l’Eau, 1997), and Germany (mainly in
what was West Germany) has more than 1100, again typically for populations
below 1000 (Bucksteeg, 1987). In New Zealand WSP are the most common
form of wastewater treatment, with 100 of the 160 plants serving populations
under 1000 being WSP (Archer and Mara, 2003*); this included the former
Mangere WSP in Auckland (see below). There are many WSP systems in
Australia, including those at the Western Treatment Plant in Melbourne. They
are common in all parts of the developing world, where they can serve large
populations, for example, the Dandora WSP near Nairobi, Kenya serve a
Waste Stabilization Ponds 95
Figure 9.5 The Phase I WSP at Dandora, Nairobi, Kenya
sewered population of ~1 million, and the Al Samra WSP near Amman, Jordan
serve a population of ~2.6 million.
Examples of large WSP systems
Dandora, Nairobi
Phase I of the Dandora WSP, which was commissioned in 1980, comprises two
series of ponds, each of which has a 21-ha primary facultative pond and three
9-ha maturation ponds (Figure 9.5); the design flow was 30,000 m3/day. Phase
II, commissioned in 1992, added a further six series, each comprising a 21-ha
facultative pond and three 4.5-ha maturation ponds; the design flow was
80,000 m3/day, and the total (Phase I and II) pond area is ~300 ha. Phase II
also included three pilot-scale anaerobic ponds ahead of one of the new series
(Chapter 10; see Pearson et al, 1996a*). Phase III will be the provision of
anaerobic ponds as the first stage in all eight series, so taking the design flow
to 160,000 m3/day – equivalent to a population of ~2 million.
Al Samra, Amman
This WSP system (see Figure 13.13) was commissioned in 1985 and currently
treats a wastewater flow of ~150,000 m3/day in three series, each comprising
two 3.2-ha anaerobic ponds, four 7.25-ha facultative ponds and four 6.25-ha
maturation ponds – a total pond area of 181 ha. Full details are given by
Al-Salem and Lumbers (1987). The design flow was 68,000 m3/day, so the
ponds are overloaded hydraulically by a factor of 2.2; even so, BOD removal
96 Domestic Wastewater Treatment in Developing Countries
is 73–85 per cent (and no electromechanical treatment plant as overloaded as
this could do so well) (Mara and Pearson, 1998*).
The arrangement of ponds in each series is not the best – all 10 ponds are
in series, and all but the last two maturation ponds are overloaded organically,
as well as hydraulically (however, the 10 ponds in each series could be
reconfigured relatively easily to provide, for example, 5 series of 6 ponds or
even 10 series of 3 ponds). The WSP discharge into the King Talal reservoir,
which has a capacity of 78 million m3. This reservoir also receives some
freshwater input, as well as some other treated wastewaters; it effectively acts
as a wastewater storage and treatment reservoir (Chapter 16). The contents of
the reservoir are used for unrestricted irrigation in the Jordan Valley (the east
bank of the River Jordan), and it is the microbiological quality of the reservoir
effluent (rather than that of the ponds) that is important for its use for
unrestricted irrigation.
Werribee, Melbourne
The Western Treatment Plant at Werribee in Melbourne, Australia commenced
operations in 1892 as a ‘sewage farm’ (Penrose, 2001*). Currently it receives
530,000 m3 of wastewater per day from a population of 1.6 million (55 per
cent of the city’s total population) and from various industries (mainly
distilleries, breweries, slaughterhouses and rendering plants, food processing
factories, tanneries, paper mills and heavy industrial laundries) which
contribute ~22 per cent of the flow and ~55 per cent of the BOD (Melbourne
Water, 2002*). WSP systems were introduced in 1937; there are now three
modern WSP systems (the ‘115E, 55E and 25W lagoons’), which together treat
69 per cent of the annual raw wastewater flow in a total pond area of 1667
ha. Each series comprises 10 ponds with an overall retention time of 50–120
days. The final effluent is discharged into Port Phillip Bay. Effluent quality is
very high: unfiltered BOD, 10–43 mg/l; suspended solids, 30–62 mg/l; total N,
24–27 mg/l; total P, ~6 mg/l; and E coli, 50–500 per 100 ml. The mean
temperature of the coldest month (July) is 13ºC and the warmest month
(January) 25°C.
The ‘55 East lagoon’ system was upgraded in 2001. It comprises 10 ponds
in series, each 200 x 1500 m (Figure 9.6). Pond 1 is a hybrid system: the first
400 m has a depth of 6 m and acts as an anaerobic pond, the first half of
which is covered and the biogas collected (Chapter 13). The rest of Pond 1 is
1.9 m deep and is aerated by 50 30 kW surface aerators. Ponds 2–4 are
maturation ponds, as are ponds 6–10. Pond 5 contains an activated sludge
plant for nitrification and denitrification, which receives 30 per cent of its
flow direct from the anaerobic pond and 70 per cent from Pond 4.
Nitrification occurs in the aerated tanks and denitrification in the anoxic
tanks. Secondary sedimentation tanks settle the activated sludge which is
returned to the anoxic tanks; excess sludge is discharged into the aerated
section of Pond 1. The clarified effluent is discharged into the remaining
section of Pond 5 (200 x 900 m), and thence into Ponds 6–10. The final
effluent is discharged into Port Phillip Bay.
Waste Stabilization Ponds 97
Note: Photograph taken during the covering of the anaerobic section of Pond 1 and prior to the
provision of the nitrification–denitrification activated sludge plant in Pond 5.
Source: Courtesy of Melbourne Water.
Figure 9.6 The 55 East WSP Series at Werribee, Melbourne, Australia
The WSP at Melbourne Water’s Western Treatment Plant at Werribee have
undergone several major modifications since their introduction in the late
1930s in response to changing environmental priorities. The recent addition of
98 Domestic Wastewater Treatment in Developing Countries
the nitrifying/denitrifying activated sludge unit was necessary as the regulator
(the Environmental Protection Agency of the State of Victoria) required a
maximum discharge into Port Phillip Bay of 3100 tonnes of total N per year
(N, rather than P, is the limiting nutrient for eutrophication of the Bay), and it
was not possible to guarantee this by conventional WSP within the land area
available. Melbourne Water has shown that an ‘enhanced lagoon system’ (of
which 55E is the largest example in the world) can meet stringent ‘21st
century’ environmental quality objectives.
Mangere, Auckland
The Mangere WSP in Auckland were commissioned in 1960 as secondary and
tertiary treatment units following primary sedimentation. The WSP, which
were built on land reclaimed from a tidal mudflat, comprised three facultative
ponds in parallel and a single maturation pond in series which discharged into
Manukau harbour at high tide (Figure 9.7). (Building the ponds in Manukau
harbour destroyed a traditional Maori shellfishing area, but in 1960 this was
not a major consideration – at least not to the local non-Maori decision
makers. Today attitudes have changed and such an approach would be
unthinkable.) The area of each facultative pond was 119–135 ha, and that of
the maturation pond 125 ha – a total of 512 ha. The effluents from the
facultative ponds were recirculated to their inlets, with a recirculation ratio of
~2.7. As the population grew, the ponds became overloaded and they were
mechanically aerated until four large biofilters (Chapter 19) were installed in
1980. The WSP then received a variable mixture of primary and secondary
effluents so that the surface BOD load on the facultative ponds could be
controlled throughout the year, from 60–80 kg/ha day in winter to 110–140
kg/ha day in summer. This was necessary to control midge (Chironomus
zealandicus) breeding and to prevent algal crashes with concomitant odour
release (the algal crashes were due to infection with the fungal parasite
Pseudosphaerita euglenae and grazing by predatory rotifers, Brachionus spp).
The algal crashes occurred once or twice each year with major odour emissions
which travelled 5–10 km and led, understandably, to complaints from local
residents. Midge control was achieved by spraying the shallow margins of the
ponds with Malathion; however, this was not always successful and swarms of
midges also affected the local residents fairly regularly.
Recently, under the Resource Management Act 1991, the local regulator
(Auckland Regional Council) set new effluent quality requirements for
wastewater discharges into Manukau harbour; in particular, in order to
prevent eutrophication (especially the proliferation of the red seaweed
Gracilliaria) and to avoid fish toxicity problems, the ammonia consent level
was reduced from ≤38 mg N/l to ≤5 mg N/l in winter and ≤3 mg N/l in summer.
The existing plant, including the WSP, was unable to achieve this, and so the
decision was taken to build a completely new land-based treatment system, to
include nutrient removal, by 2003. The WSP could have been used for
disinfection to achieve the required E coli count of ≤80 per 100 ml, but midge
breeding and algal crashes followed by odour release would have remained a
Waste Stabilization Ponds 99
Note: The maturation pond is at top right.
Source: Courtesy of Watercare Services Ltd, Auckland
Figure 9.7 The Mangere WSP, Auckland, New Zealand, in 1996
real risk (Lawty et al, 1996*). This risk was no longer acceptable and
consequently the WSP were decommissioned in 2002; their site is now once
again part of Manukau harbour (Watercare Services Ltd, 2002*). Disinfection
is now achieved by UV radiation as the final unit process of the new landbased treatment plant.
The Mangere ponds served Auckland satisfactorily for over 40 years,
despite the problems with midges and odour release. They demonstrated well
the ability of WSP to meet their discharge requirements until it was decided to
impose a very much stricter ammonia standard. Reclaiming more land from
the harbour was not an option, and so the Mangere ponds had to be
decommissioned. Over 4 million m3 of sludge were painstakingly dredged from
the ponds before the pond area was opened to the sea. Some 13 km of newly
restored coastline has been landscaped and planted with native coastal species.
The rapid recovery of the marine ecosystem in the large tidal embayment
100 Domestic Wastewater Treatment in Developing Countries
formerly occupied by the ponds has been well received by the regulator and
environmentalists, and particularly by the local Maori people who have
welcomed the return of their traditional fishing grounds.
Pearson et al (1987a) studied the WSP at Cajamarca (2675 m above mean sea
level) and Juliaca (3827 m) in the high Andes of Peru. The ponds at Cajamarca
(three series, each comprising a facultative and a maturation pond) achieved a
filtered COD (chemical oxygen demand) removal of ~85 per cent at a pond
water temperature of 15–21°C (which remained reasonably constant despite
the daily air temperature varying between 5 and 24°C, a diurnal range of
19 degC). Chlorophyll a levels (see Chapter 11) in the facultative ponds were
~1500 µg/l, and sludge accumulation was minimal: only 5 cm in the facultative
ponds after five years of operation.
The WSP at Juliaca comprised eight facultative ponds in parallel. Filtered
COD removal was ~80 per cent, and chlorophyll a levels were ~320 µg/l. Inpond temperatures were 6–9°C and the daily air temperature varied between
–5 and 14°C (a diurnal difference also of 19 degC). Sludge accumulation was
also minimal: 5.5 cm after three years.
Thus high-altitude WSP can be designed in the same way as low-altitude
ponds (Chapters 10–12), although, of course, due to lower temperatures they
are larger than low-altitude ponds (see also Juanicó et al, 2000*). Care must
be taken, however, at altitudes above 3500 m as mean in-pond temperatures
are around the same as mean air temperatures in the coldest month, rather
than being 2–4 degC warmer. Interestingly, anaerobic activity in primary
facultative ponds does not appear to be adversely affected by temperatures
that are low throughout the year, presumably because the anaerobic bacteria
have adjusted to them. In contrast, methanogenesis in low-altitude WSP
decreases in the cool season.
How well do WSP stand up against other treatment processes, such as
constructed wetlands, UASBs and aerated lagoons?
WSP vs constructed wetlands
Constructed wetlands (Chapter 17) are secondary treatment units – that is they
have to be preceded by a septic tank or an anaerobic pond to remove the solids
that would otherwise clog the gravel bed of the wetland. As shown in the
Design Example in Chapter 17, constructed wetlands require more land than a
secondary facultative pond and a large amount of gravel. Moreover, Tanner
(2001*) has demonstrated that the plants used in gravel-bed constructed
Waste Stabilization Ponds 101
wetlands (such as the common reed, Phragmites australis) are unnecessary for
the removal of BOD, COD, suspended solids, phosphorus and E coli (see also
Ayaz and Akça, 2001*; Regmi et al, 2003). This implies that the gravel bed
itself, rather than the plants growing in it, is responsible for these removals
and that a rock filter (Chapter 12) would therefore be more appropriate.
Tanner did find, however, that the plants were needed for nitrogen removal;
this suggests that constructed wetlands should only be considered for use when
the regulator has specified a maximum effluent ammonia concentration (and
they can be designed specifically for ammonia removal using equation 17.5).
The question that designers must answer is: can WSP achieve the required
ammonia level (equations 12.12–12.18) at lower cost than constructed
Upflow anaerobic sludge blanket reactors (UASBs) (Chapter 18) are very
efficient anaerobic reactors with short retention times (~8 hours). However,
they are not significantly more efficient than anaerobic ponds, especially highrate anaerobic ponds (Chapter 10), but they are much more expensive to
construct, and overall the land area savings are small (see the Design Example
in Chapter 18).
WSP vs aerated lagoons
Aerated lagoons (Chapter 20) are activated sludge units operated without
sludge return. Retention times are 2–6 days, with 4 days being a typical value.
The question to be answered is: is a 4-day aerated lagoon cheaper than a 1day anaerobic pond and a 4-day secondary facultative pond? The answer to
this question must include the annual cost of the electricity used to power the
aerators. How much extra land could be purchased with this annual
expenditure on electricity?
Macrophytes are higher plants (whereas ‘microphytes’ are algae); they are
occasionally, but generally inappropriately, used in or on WSP for algal and/or
nutrient removal.
Floating macrophyte ponds
Macrophytes such as water hyacinth (Eichhornia crassipes) or water lettuce
(Pistia stratiotes) are grown over the whole surface of a maturation pond in
order to reduce algal suspended solids in the final effluent by depriving the
algae of light. The roots of the plants also help to reduce nitrogen and
phosphorus levels. However, they have the major disadvantage of encouraging
mosquito breeding; this can be controlled by the introduction and maintenance
102 Domestic Wastewater Treatment in Developing Countries
of mosquito fish, such as Gambusia, which eat mosquito larvae, but this is an
extra O&M task – making sure the fish stay alive. Moreover, the plants need
regular harvesting every few months, otherwise they die in the pond and lead
to a deterioration in effluent quality. As algae are vital for faecal bacterial and
viral removal in WSP (Chapter 12), floating macrophyte ponds should only be
used (if used at all) when E coli numbers have been reduced to the required
level in the preceding unplanted maturation ponds.
Duckweeds of the family Lemnaceae are being increasingly recommended
as a floating macrophyte for ponds (eg Skillicorn et al, 1993*; Iqbal, 1999*),
since they are high in protein and also β-carotene (the pigment that gives egg
yolks a deep gold colour). Harvested duckweeds can be used to feed certain
species of carp and also hens (to ensure their eggs have golden yolks). However,
they are troublesome to grow as they are so small (only a few mm) that they
are easily blown by the wind into a corner of the pond unless wind action is
minimized by placing lengths of bamboo across the pond surface. Fish yields
are no higher than those from algal fishponds (Chapter 22), and golden-yolked
eggs may not be as high a consumer preference in developing countries as it is
in industrialized countries. Furthermore, nutrient removal in duckweed ponds
is no higher than that in maturation ponds (Chapter 12) (see Zimmo, 2003).
Rooted macrophyte ponds
Rooted macrophytes of the type used in constructed wetlands (Chapter 17)
can be planted in maturation ponds, which therefore become free-watersurface constructed wetlands. In France it used to be common practice to plant
50–100 per cent of the final maturation pond with the common reed
(Phragmites australis), but this led to such massive mosquito breeding
(especially of Coquillittidia spp and Mansonia spp, the larvae of which obtain
their oxygen from the plants’ lacunae, which transport oxygen from the leaves
to the roots) that the practice had to be discontinued (Ringuelet, 1983).
There are two types of advanced pond systems: high-rate algal ponds, which
were originally designed for algal production, and the ‘Advanced Integrated
Pond’ system. They are included here for completeness, but they are not
applicable in developing countries (although increasingly advocated as such).
High-rate algal ponds
High-rate algal ponds (HRAP) were devised by Professor William Oswald at
the University of California at Berkeley in the 1960s (see Oswald, 1988a).
They receive settled wastewater and they were originally designed to maximize
algal production, rather than to optimize wastewater treatment, simply
because algae are 50–60 per cent protein and protein yields from HRAP are so
much greater than from conventional agriculture (~30,000 kg protein/ha year,
Waste Stabilization Ponds 103
compared with <1000 kg/ha year from soya beans), and algae are cheap to
produce – around US$0.25/kg dry weight (Oswald, 1988b). In a food-short,
especially protein-short, world HRAP are immediately attractive: even if the
algae harvested from HRAP are not used for direct human consumption, they
can be used as supplementary animal (commonly, chicken) feeds (but health
food shops in industrialized countries have algal tablets on sale; they are held
to be useful for dieting ‘the natural way’, and there is even a book called
Chlorella: The Emerald Food – Bewicke and Potter, 1984).
The HRAP is arranged in a ‘race track’ configuration and its contents are
gently stirred by electrically driven paddle wheels to prevent the algae from
settling (the layout is essentially the same as that used for oxidation ditches –
Figure 20.3), with a depth of 20–50 cm and a retention time of 2–6 days. Algal
production is very high, around 15 g dry weight/m2 day, equivalent to ~30 t
protein/ha/year (see Oswald, 1995). The next two processes are algal
harvesting and drying. The best way to harvest the algae from the HRAP
effluent is dissolved air flotation. Drying can be done in the sun or, if the HRAP
is preceded by a conventional primary sedimentation tank (rather than an
anaerobic pond), the methane generated in an anaerobic digester treating the
primary sludge can be burnt to heat air to dry the algae.
It will be apparent that the use of HRAP for maximal algal production is a
complicated process, at least from the perspective of a wastewater treatment
engineer (but not, perhaps, from that of an industrial microbiologist).
However, the algae grown at high densities in HRAP are susceptible to fungal
infections which can cause the population to ‘crash’. In short, and despite their
advantage of high protein yields, HRAP are too complicated a process for
wastewater treatment in developing countries. There are very few full-scale
HRAP even in industrialized countries, which suggests they might be just too
complicated – full stop. Moreover, once the algae have been harvested and
dried, they must be packaged and sold – otherwise there is no point in
producing them in the first place. However, few, if any, wastewater treatment
engineers in either industrialized or developing countries are likely to have the
required marketing skills.
Advanced integrated pond systems
The ‘Advanced Integrated Wastewater Ponding System’ (AIWPS, which is a
US registered trademark, so it should be written AIWPS®) was developed in
California as a successor to HRAP (Oswald, 1991), presumably because of the
difficulties encountered in practice with commercial algal production in HRAP.
AIWPS is primarily a wastewater treatment process, and algal production is
only a secondary consideration. It is also termed an ‘Advanced Integrated Pond
System’ (Department of Energy, 1993). The AIWPS/AIPS consists of four or
five ponds in series, as follows:
a 4–5 m deep ‘advanced’ facultative pond which contains a solids ‘digester
pit’; the raw wastewater, after preliminary treatment (Chapter 8) is
104 Domestic Wastewater Treatment in Developing Countries
introduced at the base of the digester pit, which acts like an anaerobic
pond within the facultative pond;
an HRAP, part of the effluent from which is recirculated to the facultative
pond to help keep its upper layer aerobic;
an algal settling pond (the settled algae can be removed and used as
fertilizer); and
one or two maturation ponds for disinfection (optional).
Oswald (1991) reports the following retention times for the Advanced
Integrated Pond System at St Helena, California: advanced facultative pond,
20 days; HRAP, 10 days; and algal settling pond, 5 days – a total of 35 days.
For the AIPS at Hollister, California, the retention times are: advanced
facultative pond 32 days; HRAP, 10 days; and algal settling pond, 7 days – a
total of 49 days.
Advanced pond systems
‘Advanced pond systems’ are beginning to be used in New Zealand (a
‘technology transfer’ from California) (Craggs et al, 2003*). They have the
same types of ponds as the Californian AIWPS/AIPS, but with different
retention times, for example: HRAP, 7.5 days; algal settling pond, 2 days;
maturation pond, 20 day. Two algal settling ponds (each with a 2-day retention
time) are provided in parallel: one is operational while algae are removed from
the other.
It appears, therefore, that these ‘advanced’ or ‘advanced integrated’ pond
systems have no advantages over modern waste stabilization ponds designed
as described in Chapters 10–12 or, in water-short regions, over the wastewater
storage and treatment reservoirs (including the hybrid WSP–WSTR system)
described in Chapter 16. (Of course, if an industrial microbiologist/
entrepreneur wishes to produce algae commercially, then these advanced pond
systems could be very advantageous.)
Anaerobic Ponds
Anaerobic ponds (Figure 10.1) are usually the first type of pond used in a series
of ponds (Chapter 9). They are 2–5 m deep and receive such a high organic
loading (usually >100 g BOD/m3 day, equivalent to >3000 kg/ha day for a
depth of 3 m) that they contain no dissolved oxygen and no algae, although
occasionally a thin film of Chlamydomonas may be present at the surface.
They function much like open septic tanks, and their primary function is BOD
(biochemical oxygen demand) removal. Anaerobic ponds work extremely well:
a properly designed and not significantly underloaded anaerobic pond will
achieve >60 per cent BOD removal at 20°C. Retention times are short: for
wastewater with a BOD of ≤300 mg/l, for example, 1 day is sufficient at a
temperature of 20°C.
BOD removal is achieved by the sedimentation of settleable solids and their
subsequent anaerobic digestion in the resulting sludge layer; this is particularly
intense at temperatures above 15°C when the pond surface literally bubbles
with the release of biogas (around 70 per cent methane and 30 per cent carbon
dioxide). The bacterial groups involved are the same as those in any anaerobic
reactor (Chapter 3), and those in anaerobic ponds are equally sensitive to the
same toxicants, one of the most important of which is a pH below ~6.2.
Many compounds in industrial wastewaters are toxic to algae, and
treatment in anaerobic ponds prior to facultative and maturation ponds
(Chapters 11 and 12) is necessary to avoid this. Heavy metals are precipitated
as metal sulphides, and many organic toxicants (eg phenol) are degraded to
non-toxic forms. Floating materials (including oils) and scum, which block out
the light needed in facultative ponds for algal photosynthesis, are retained in
anaerobic ponds.
There is a gradual accumulation of digested solids in anaerobic ponds, so
regular desludging once every 1–3 years is required (Chapter 14). Scum
accumulates on the surface, but does not need to be removed (it helps keep the
pond anaerobic), unless fly breeding reaches nuisance level (Chapter 14).
106 Domestic Wastewater Treatment in Developing Countries
Figure 10.1 Anaerobic Pond, with Partial Scum Coverage, at Ginebra,
Valle del Cauca, Southwest Colombia
Designers have in the past been afraid to incorporate anaerobic ponds in case
they cause odour. Hydrogen sulphide, formed mainly by the anaerobic
reduction of sulphate by sulphate-reducing bacteria such as Desulfovibrio spp,
is the principal potential source of odour. However, in aqueous solution
hydrogen sulphide is present as either dissolved hydrogen sulphide gas (H2S)
or the bisulphide ion (HS–), with the sulphide ion (S2–) only being formed in
significant quantities at high pH (Figure 10.2). At the pH values normally
found in well designed anaerobic ponds (around 7.5), most of the sulphide is
present as the odourless bisulphide ion. Odour is only caused by escaping
hydrogen sulphide molecules as they seek to achieve a partial pressure in the
air above the pond which is in equilibrium with their concentration in it
(Henry’s law). Thus, for any given total sulphide concentration, the greater the
proportion of sulphide present as HS–, the lower the release of H2S. Odour is
not a problem if the recommended design loadings (Table 10.1) are not
exceeded and if the sulphate concentration in the raw wastewater is <500 mg
SO42– /l (Gloyna and Espino, 1969). Sulphate concentrations are usually much
less than this; they depend on the amount of sulphate in the local drinking
water (for which the World Health Organization, 1993, recommends a limit
of 250 mg/l as SO42– ), and on the local usage of detergents which contain large
amounts (up to 40 per cent by volume) of sodium sulphate as a ‘bulking’ agent.
Anaerobic Ponds 107
It is always worth checking the local drinking water – in parts of southern
Spain, for example, the drinking water contains 600–1200 mg SO2–
4 /l and
odour was a serious problem with anaerobic ponds until the sulphate
concentration was reduced to <250 mg/l at the local water treatment plant. A
further point to consider in coastal tourist areas is whether hotels use seawater
for toilet flushing – seawater contains around 3000 mg SO42– /l – so if the hotel
wastewaters are treated in anaerobic and facultative ponds, there will be
serious odour problems (Frederick-van Genderen, 1995).
Paing (2001) presents a model for odour release from anaerobic ponds in
southern France receiving domestic wastewater with 100–220 mg SO2–
4 /l; inpond temperatures were 8–26°C. It was found that sulphide was released to
the atmosphere at rates of 25–710 mg S per m2 of anaerobic pond surface area
per day, which resulted in H2S concentrations in the air of 0.3–8.5 mg S/m3
and complaints from local residents.
A small amount of sulphide is beneficial as it reacts with heavy metals to
form insoluble metal sulphides which precipitate out, but concentrations of
50–150 mg/l can inhibit methanogenesis (Pfeffer, 1970). A further important
advantage of small concentrations (>3 mg/l) of sulphide in anaerobic ponds is
that they are rapidly lethal to Vibrio cholerae (Oragui et al, 1993; Arridge
et al, 1995*).
Designs now exist for covering anaerobic ponds to facilitate methane
recovery (DeGarie et al, 2000*; see also Chapter 13 and Figure 16.3). This
also provides additional security against odour release since, even if the
methane is not to be used for power generation, the gases can be flared off.
Source: Sawyer et al (2002)
Figure 10.2 Variation of the Proportions of Hydrogen Sulphide, Bisulphide
and Sulphide with pH in Aqueous Solutions
108 Domestic Wastewater Treatment in Developing Countries
Anaerobic ponds can be satisfactorily designed, without the risk of odour
release, on the basis of volumetric BOD loading (λv, g/m3 d), which is given
λv = LiQ/Va
where Li is the influent BOD, mg/l (= g/m3), Q is the flow, m3d; and Va is the
anaerobic pond volume, m3.
The permissible design value of λv increases with temperature, but there
are too few reliable data to enable the development of a suitable design
equation. Mara and Pearson (1998*) recommend the design values given in
Table 10.1; these recommendations were based on those of Meiring et al
(1968) that λv should lie between 100 and 400 g/m3 day, the former in order
to maintain anaerobic conditions and the latter to avoid odour release.
However, in Table 10.1 the upper limit for design is set at 350 g/m3 day in
order to provide an adequate margin of safety with respect to odour.
Once a value of λv has been selected, the anaerobic pond volume is then
calculated from equation 10.1. The mean hydraulic retention time in the pond
(θa, d) is determined from:
θa = Va/Q
Retention times in anaerobic ponds greater than 1 day should not be used. If
equation 10.2 gives a value of θa <1 day, a value of 1 day should be used. The
corresponding value of Va is Q m3.
Anaerobic ponds are usually 2–5 m deep, with a value of 3 m usually being
assumed for process design (and adjusted, if necessary, during the physical
design stage – Chapter 13). The anaerobic pond area (Aa, m2) is then given by:
Aa = Qθa/Da = LiQ/λvDa
where Da is the anaerobic pond working (ie liquid) depth, m.
The performance of anaerobic ponds increases significantly with
temperature, and the design assumptions for BOD removal (needed for the
design of the receiving facultative pond) given in Table 10.1 can be confidently
adopted. These are based on experience with anaerobic ponds in Germany in
winter (T <10°C) (Bucksteeg, 1987), and in northeast Brazil at 25°C where
little variation in BOD removal was found at retention times of 0.8–6.8 days
(Table 10.2). Pearson et al (1996a*) also reported little variation of BOD
removal with retention time at the Dandora ponds serving Nairobi, Kenya:
79–86 per cent removal at retention times of 2.5–9.5 days; the removal of
Anaerobic Ponds 109
Table 10.1 Design Values of Volumetric BOD Loadings on and Percentage
BOD Removals in Anaerobic Ponds at Various Temperatures
Temperature (ºC)
Volumetric loading (g/m3 day)
BOD removal (%)
20T – 100
10T + 100
2T + 20
2T + 20
T = temperature,°C
79 per cent was achieved in the coldest month, July, when the mean monthly
air temperature is 17°C – that is the removal was very much higher than the
56 per cent expected for 17°C from Table 10.1.
Design temperature
The design temperature for anaerobic ponds (and facultative ponds – Chapter
11) is usually taken as the mean air temperature of the coldest month. This is
slightly conservative as the pond temperature is 1–2 degC higher than the air
temperature in the coldest month.
The mean monthly temperature is the mean of the mean daily temperatures
– that is the 28–31 day average of the mean of the daily maximum temperature
and the daily minimum temperature (these are measured every day at
meteorological stations, usually at 8 am).
Sludge accumulation
As noted in Chapter 13, the rate of accumulation of digested sludge in
anaerobic ponds in warm climates is usually taken as 0.04 m3/person year.
This is, however, only a rough estimate, and in northeast Brazil much lower
rates were measured: of the order of 0.01 m3/person year (Silva, 1982). Initial
rates of sludge accumulation are higher as the sludge volume comprises two
zones: one for sludge digestion and one for the storage of digested sludge –
initially the volume of the former is a higher proportion of the total, so
Table 10.2 Variation of BOD Removal with BOD Loading and Retention
Time in Anaerobic Ponds in Northeast Brazil at 25 ºC
Retention Time (days)
Source: Silva (1982)
Volumetric BOD loading (g/m3 d)
BOD removal (%)
110 Domestic Wastewater Treatment in Developing Countries
accumulation rates are higher; as time passes its proportion of the total
declines and accumulation rates become lower. Nelson (2002) describes a
mechanistic model for sludge accumulation in primary facultative ponds; the
model includes sedimentation of influent settleable solids, digestion of the
settled solids and sludge compaction (a similar approach is taken by Metcalf
and Eddy, Inc (1991) for the design of sedimentation ponds following aerated
lagoons – Chapter 20). The management of sludge from anaerobic ponds is
fully described by Franci (1999).
Drying beds
Sludge removed from anaerobic ponds can be conveniently dewatered in sludge
drying beds. Moisture loss is by evaporation and percolation through the sandgravel bed which comprises the base of the drying bed (300 mm of 0.3–0.75
mm sand over 300 mm of 5–25 mm gravel). The drying bed area should be
~0.025 m2/person, and the sludge should be applied at the start of the dry
season (Chapter 13).
High-rate anaerobic ponds are a recent development by Professor Miguel Peña
Varón at the Universidad del Valle in Cali, Colombia (Peña Varón, 2002*;
Peña Varón et al, 2002*). They combine the higher performance of upflow
anaerobic sludge blanket reactors (UASBs – Chapter 18), which can achieve
70 per cent BOD removal at a retention time of only 6 hours, with the
constructional and operational simplicity of anaerobic ponds. Several designs
were evaluated by computer modelling, and some were then tested on a pilot
scale. The best design, shown in Figure 10.3, introduces the wastewater at the
base of the ‘mixing pit’ (as in a UASB) and separates solids digestion and solids
settlement (as a UASB does); it achieved unfiltered and filtered COD (chemical
oxygen demand) removals of 79 and 78 per cent, respectively, at a retention
time of 0.6 day and a COD loading of 1115 g/m3 day (equivalent to ~450 g
BOD/m3 day). A conventional (ie low-rate) anaerobic pond, operating at the
same retention time and loading (and therefore overloaded) achieved unfiltered
and filtered COD removals of only 15 and 22 per cent, respectively – thus
clearly demonstrating the ability of the high-rate anaerobic pond to accept
high loads and achieve very high treatment efficiency.
With normal strength domestic wastewater it is not advantageous to have two
(or more) anaerobic ponds in series. Silva (1982) found that a 0.8-day
anaerobic pond reduced the BOD from 240 mg/l to 60 mg/l at 25°C in
northeast Brazil, but a second 0.4-day anaerobic pond in series reduced it to
only 46 mg/l. However, for the treatment of much stronger agro-industrial
Anaerobic Ponds 111
Plastic mesh
Mixing pit
H = 3.2 m
h = 1.7 m
a = 2.3 m
b = 4.9 m
Source: Peña Varón (2002*)
Figure 10.3 High-rate Anaerobic Pond with a Mixing Pit
112 Domestic Wastewater Treatment in Developing Countries
wastewaters, a series of anaerobic ponds is very efficient. McGarry and Pescod
(1970) used a series of five anaerobic ponds for the treatment of tapioca starch
wastewater in Thailand; each pond was loaded at 224 g BOD/m3 day and the
BOD was reduced from 3800 mg/l to 224 mg/l. Rao and Viraraghavan (1985)
used two anaerobic ponds in series to treat distillery wastewater in southern
India: the first pond reduced the BOD from 40,000 mg/l to 5000 mg/l and the
second reduced it to 2000 mg/l (further treatment was provided in an
oxidation ditch – Chapter 20, presumably because there was no space for more
anaerobic ponds).
Design an anaerobic pond to treat the following wastewater:
Q = 10,000 m3/day (ie 100,000 persons each producing 100 litres of
wastewater per day),
Li = 300 mg/l,
T = 25°C.
From Table 10.1, λv = 350 g/m3 day for 25°C. Assume a depth of 3 m; then
from equations 10.2 and 10.3:
Aa = LiQ/λvDa
= (300 x 10,000)/(350 x 3)
= 2860 m2
θa = AaDa/Q
= (2860 x 3)/10,000
= 0.86 day
This is <1 day, the minimum retention time in anaerobic ponds. Therefore
recalculate Aa from equation 10.3:
Aa = Qθa/Da
= 10,000 x 1/3
= 3340 m2
Anaerobic Ponds 113
From Table 10.2, the BOD removal is 70 per cent. Thus the effluent BOD is
(0.3 x 300) = 90 mg/l. The anaerobic pond effluent requires further treatment
in a secondary facultative pond (Chapter 11), a wastewater storage and
treatment reservoir (Chapter 16) or a horizontal-flow constructed wetland
(Chapter 17). If sufficient land for these natural wastewater treatment
processes is not available, consideration should be given to treating the
anaerobic pond effluent in a biofilter (Chapter 19), an aerated lagoon or an
oxidation ditch (Chapter 20).
Facultative Ponds
Facultative ponds are of two types: primary facultative ponds which receive
raw wastewater (after preliminary treatment – Chapter 8 and see Figure 9.1),
and secondary facultative ponds which receive settled wastewater (usually the
effluent from anaerobic ponds). They are designed for BOD (biochemical
oxygen demand) removal on the basis of a relatively low surface BOD loading
in the range 100–400 kg/ha day (see equation 11.1 below) to permit the
development of a healthy algal population, as the oxygen for BOD removal by
the pond bacteria is mostly generated by algal photosynthesis. Facultative
ponds are coloured dark green as a result of the large numbers of micro-algae
in them, although they may occasionally appear red or pink (especially when
slightly overloaded) due to the presence of anaerobic purple sulphide-oxidizing
photosynthetic bacteria (see ‘Purple ponds’ later in this chapter). The algae
that tend to predominate in the turbid waters of facultative ponds (Table 11.1)
belong to motile genera (such as Chlamydomonas, Pyrobotrys and Euglena)
as these can optimize their vertical position in the pond water column in
relation to incident light intensity more easily than non-motile forms (such as
Chlorella, although this is also common in facultative ponds). The
concentration of algae in a facultative pond depends on loading and
temperature (see Figure 11.5 below), but is usually in the range 500–2000 µg
chlorophyll a per litre.
As a result of the photosynthetic activities of the pond algae, there is a
diurnal variation in the concentration of dissolved oxygen (Figure 11.1). After
sunrise, the dissolved oxygen level gradually rises, in response to
photosynthetic activity, to a maximum in the mid-afternoon, after which it
falls to a minimum during the night when photosynthesis ceases and algal (as
well as bacterial) respiratory activity consumes oxygen. The position of the
‘oxypause’ (the depth at which the dissolved oxygen concentration reaches
zero) similarly changes, as does the pH, since at peak algal activity carbonate
and bicarbonate ions react to provide more carbon dioxide for the algae, so
leaving an excess of hydroxyl ions with the result that the pH can rise to >9.4,
which rapidly kills most faecal bacteria (Chapter 12).
Facultative Ponds 115
DO concentration (mg/l)
Time (h)
Figure 11.1 Diurnal Variation of Dissolved Oxygen in a Facultative Pond:
ɂ top 200 mm of pond; ɀ, 800 mm below surface
Mixing and stratification
Wind, heat and pond inlet design are the three factors of major importance
that influence the degree of mixing that occurs within a pond. Mixing
minimizes hydraulic short-circuiting and the formation of stagnant regions
(‘dead spaces’) and it ensures a reasonably uniform vertical distribution of
BOD, algae and oxygen. Mixing is the only means by which non-motile algae
can be carried up into the zone of effective light penetration (the ‘photic’ zone);
since the photic zone comprises only the top ~300 mm of the pond, much of
the pond contents would remain in photosynthetic darkness if mixing did not
occur. Mixing is also greatly influenced by the design of the inlet structure;
this is discussed in greater detail in Chapter 13.
The depth to which wind-induced mixing is felt is largely determined by
the horizontal distance over which the wind is in contact with the water (the
‘fetch’): an unobstructed contact length of about 100 m is required for
maximum mixing by wind action. The importance of wind action is clearly
demonstrated by the following, admirably simple, experiment which was
conducted on a facultative pond in Zambia: a 2-m high fence with no openings
was erected around the pond and within a few days it turned anaerobic; when
the fence was removed aerobic conditions were rapidly re-established (Marais,
116 Domestic Wastewater Treatment in Developing Countries
Table 11.1 Examples of Algal Genera Found in Facultative and
Maturation Ponds
Facultative ponds
Maturation ponds
a An identification key to pond algae is given in Mara (1997*) and Mara and Pearson (1998*)
b *, motile; +, present; –, absent
c The Cyanophyta (blue green algae) are strictly the cyanobacteria (blue green bacteria)
In the absence of mixing thermal stratification quickly occurs. The warm upper
layers are separated from the colder lower layers by a thin static region of
abrupt temperature change known as the ‘thermocline’. Non-motile algae
settle through the thermocline into photosynthetic darkness where they are
unable to produce any oxygen; instead they exert an oxygen demand, with the
result that conditions below the thermocline rapidly become anoxic. Above
the thermocline the motile algae move away from the hot surface waters
(which may have a temperature >30°C), and they usually form a dense layer
about 300–500 mm below the surface. This algal band is an effective light
barrier and the thermocline is usually just below the algae.
Facultative Ponds 117
The diurnal mixing pattern in a 1.5 m deep facultative pond in Lusaka,
Zambia, was thoroughly investigated by Marais (1966) and is typical of
tropical and subtropical ponds:
In the morning, if there is any wind, there is a period of complete mixing
in which the temperature is uniform throughout the pond but, owing to
the absorption of solar radiation, gradually increases.
At some time, usually during a short lull in the wind, stratification develops
abruptly and a thermocline forms. The temperature above the thermocline
increases to a maximum and then decreases, while below the thermocline
the temperature rapidly falls to a value approximately that of the earth
and thereafter remains practically constant. A certain amount of mixing
may take place above the thermocline.
In the afternoon and evening, a second period of mixing may be initiated
as follows:
(a) above the thermocline, under quiescent wind conditions, the top layers
lose their heat more rapidly than the bottom layers. The cooler top
layers sink, inducing mixing with the result that the temperature down
to the thermocline remains approximately uniform but gradually
decreases. The thermocline slowly sinks and, should the temperature
above and below it become equal due to further cooling, mixing is
initiated and sustained throughout the pond.
(b) under windy conditions, usually during periods of decreasing
temperatures, the energy imparted by the wind to the water above the
thermocline at some stage overcomes the stratification forces and
progressively mixes the warmer and colder layers adjacent to the
thermocline, causing it to be displaced downwards until the
temperature is uniform throughout and the whole pond is in a state of
This diurnal pattern of mixing and stratification also occurs in shallow lakes
in the tropics, for example in Lake George in Uganda (Viner and Smith, 1973;
see also Talling, 2001*). The same pattern occurs, but on an annual basis, in
deep lakes in temperate regions.
Thermal stratification induces algal banding and this in turn, together with
light attenuation, induces physicochemical stratification (see also Gu and
Stefan, 1995*; Kayombo et al, 2002*), especially of pH which can rise to 9–10
and rapidly cause faecal bacterial die-off (Chapter 12). However, provided the
effluent take-off point is below the maximum depth of algal banding (Chapter
13), stratification can minimize the daily effluent BOD and suspended solids
concentrations as these are largely associated with the algae in the pond
effluent (see below under the section ‘Diurnal variation in effluent quality’).
This increase in BOD and suspended solids removal may outweigh the
disadvantages of any short-circuiting induced by stratification.
118 Domestic Wastewater Treatment in Developing Countries
Sludge layer
Primary facultative ponds fulfil the functions of both anaerobic and secondary
facultative ponds, and thus the settleable solids in the raw wastewater settle in
primary facultative ponds to form a sludge layer. These solids are digested
anaerobically with the evolution of biogas. Eventually the digested solids (the
‘sludge’) have to be removed, but this is necessary only very infrequently: in
northern France primary facultative ponds are desludged once every ten years
(see Figure 14.2); in warmer climates sludge accumulation is much lower due
to more complete and more rapid digestion. Secondary facultative ponds (and
maturation ponds – Chapter 12) should not require desludging during their
design life.
Conventional design
Facultative ponds are best designed on the basis of surface BOD loading (λs.
kg/ha day), which is given by:
λs =
where Af is the facultative pond area, m2. The factor 10 arises from the units
used: LiQ is the mass of BOD entering the pond, g/day; so 10–3LiQ is in kg/day
and the area in ha is 10–4Af.
Surface loading is used for facultative ponds, rather than volumetric
loading (as for anaerobic ponds – Chapter 10), because the light needed for
algal photosynthesis arrives from the sun at the pond’s surface. Thus algal
oxygen production is a function of area, so the BOD loading (which is an
oxygen demand) must also be a function of area.
The permissible design value of λs increases with temperature (T,°C) which
is essentially a proxy for climate. The earliest relationship between λs and T is
that given by McGarry and Pescod (1970), but their value of λs is the
maximum that can be applied to a facultative pond before it fails (that is,
becomes anaerobic). Their relationship, which is therefore an envelope of
failure, is:
λs = 60 (1.099)T
Equation 11.2 cannot be used for design as it does not incorporate a factor of
safety. Mara (1987) gives the following global design equation:
Facultative Ponds 119
BOD surface loading (λs, kg/ha day)
λs = 60(1.099)T
λs = 350(1.107–0.002T)T–25
Temperature (T, °C)
Figure 11.2 Variation of Surface BOD Loading on Facultative Ponds with
Temperature According to Equations 11.2 and 11.3
λs = 350(1.107 – 0.002T)T–25
Equation 11.3 is based on a loading of 80 kg/ha day at ≤8°C in European
winters, a loading of 350 kg/ha day at 25°C in northeast Brazil, and an
arbitrary loading of 500 kg/ha day at 35°C. Note that the reference
temperature in equation 11.3 is 25°C. Equations 11.2 and 11.3 are shown
graphically in Figure 11.2.
Once a suitable value of λs has been selected, the pond area is calculated
from equation 11.1 and its retention time (θf, days) from:
θf = AfD/Qm
where D is the pond depth, m (usually 1.5 m – but see below); and Qm is the
mean flow, m3/day. The mean flow is the mean of the influent and effluent
flows (Qi and Qe), the latter being the former less net evaporation and seepage.
Thus equation 11.4 becomes:
θf = AfD/[0.5(Qi + Qe)]
120 Domestic Wastewater Treatment in Developing Countries
If seepage is negligible Qe is given by:
Qe = Qi – 0.001eAf
where e is the net evaporation rate, mm/day (meteorological stations usually
record evaporation and rainfall in mm per month, so division by 28–31 days
is necessary to obtain mm/day). Thus equation 11.5 becomes:
θf = 2AfD/(2Qi – 0.001eAf)
A minimum value of θf of 5 days should be adopted for temperatures <20°C,
and 4 days for temperatures ≥20°C. This is to minimize hydraulic shortcircuiting and to give the algae sufficient time to multiply (ie to prevent algal
Facultative pond working (ie liquid) depths are usually in the range 1–1.8 m,
with 1.5 m being the most commonly used. Depths <1 m do not prevent the
emergence of vegetation. This must be avoided, as otherwise the pond becomes
an ideal breeding ground for mosquitoes and midges. With depths >1.8 m the
oxypause is too near the surface with the result that the pond is predominantly
anaerobic rather than predominantly aerobic. This is undesirable, as the pond
would have an unacceptably low factor of safety in normal operation and so
be less able to cope with a fluctuating load or a sudden slug of heavy pollution.
In arid climates evaporation rates are high and water losses can be minimized
by increasing the depth to 2 m (even 2.5 m) and so reducing the surface area.
In colder climates (eg at high altitude) similar depths can be used in order to
preserve as much of the thermal energy of the influent wastewater as possible.
These considerations are usually more important in these extremes of climate
than those concerned with the position of the oxypause.
BOD removal
The BOD of the facultative pond effluent can be estimated from equation 5.7:
Le =
1 + k1θf
where Li is the BOD of either the raw wastewater in the case of primary
facultative ponds, or the anaerobic pond effluent (Chapter 10) in the case of
secondary facultative ponds, mg/l; and k1 is the first-order rate constant for
BOD removal, day–1, given by equation 5.8:
Facultative Ponds 121
k1(T) = k1(20)(1.05)T–20
Design values for k1(20) are 0.3 day–1 for primary facultative ponds and
0.1 day–1 for secondary facultative ponds (the value for the latter is much lower
than that for the former as essentially all the BOD removal by sedimentation
occurs in the preceding anaerobic pond).
The term Le is the unfiltered BOD which includes the BOD of the algae
present in the facultative pond effluent. This ‘algal BOD’ accounts for ~70–90
per cent of the total (ie unfiltered) BOD of the effluent. Thus the relationship
between filtered and unfiltered BOD (ie non-algal and total BOD) is:
Le(filtered) = Fna[Le(unfiltered)]
where Fna is the non-algal fraction of the total BOD (around 0.1–0.3, with a
usual design value of 0.3). As noted in Chapter 4, a pond effluent with a
filtered BOD of ≤25 mg/l is acceptable for surface water discharge in the
European Union (and efforts should be made to persuade regulators in
developing countries that this is also acceptable in their jurisdictions).
Diurnal variation in effluent quality
Due to the motility of many of the facultative pond algae, they are able to
optimize their position in the pond water column with respect to incident light
intensity and, when the pond is stratified, the algae form a band 10–20 cm
thick in the top 50 cm of the pond. As a result of this algal movement in the
pond, the quality of the effluent (which is drawn off from the pond at a fixed
depth – Chapter 13) varies throughout the 24-hour day. Figure 11.3 shows
this variation in effluent quality for a primary facultative pond in northeast
Brazil that was receiving a BOD loading of 320 kg/ha day at 25°C. Effluent
chlorophyll a concentrations varied between a few hundred to >10,000 µg/l,
and changes in chlorophyll a were mirrored by corresponding changes in the
concentrations of algae-associated parameters such as dissolved oxygen, pH,
total phosphorus, BOD and suspended solids. (The significance of dissolved
oxygen and pH changes in relation to faecal bacterial die-off is discussed in
Chapter 12.)
These large variations in effluent quality parameters mean that a single
‘grab’ sample of effluent can only give information on effluent quality at the
time the sample was collected. In general, therefore, 24-hour flow-weighted
composite samples are required for most parameters to determine the mean
daily effluent quality (Chapter 15).
Design based on uncertainty analysis
A significant contribution to facultative pond design, and to WSP design in
general, was made by von Sperling (1996b*), who recognized that there is
uncertainty about the values used for the design parameters, for example, we
122 Domestic Wastewater Treatment in Developing Countries
Total and soluble P (mg/l)
Temperature (°C)
Ammonia (mg/l)
Dissolved oxygen (mg/l)
Suspended solids (mg/l)
BOD5 (mg/l)
Time of day (hours)
Note: ᔥ, dissolved oxygen; , pH; ɀ, temperature; ᔢ, ammonia; , total and ◊, soluble
phosphorus; ɀ, BOD; ɂ, suspended solids; , chlorophyll a; and , faecal coliforms.
Figure 11.3 Diurnal Variation in Facultative Pond Effluent Quality
Faecal coliforms (log cells/100ml)
Chlorophyl a (µg/l)
Facultative Ponds 123
may design for some future population P which has a wastewater generation
of q l/person/day and a BOD contribution of B g/person day, and we then
calculate the wastewater flow and BOD as (10–3Pq) m3/day and (103B/q) mg/l,
respectively. However, even though the values of P, q and B used in any one
design are the design engineer’s best estimates, they are in effect being used as
‘certain’ values which do not admit any variation. In practice, as recognized
by von Sperling, there is some degree of uncertainty in their values; in other
words, instead of taking P as, for example, 100,000, P is be allowed to be
anywhere between, say, 80,000 and 120,000 – that is, there is an uncertainty
in the value of P of, in this case, ± 20 per cent. This principle of uncertainty is
applied to the values of all the parameters used for (in this case) facultative
pond design. In equations 11.1, 11.3, 11.7, 5.7, 5.8 and 11.8, there are the
following nine independent parameters used for facultative pond design:
Population, P
BOD contribution, B (g/person day)
Wastewater flow, q (l/person day)
Temperature, T (ºC)
Pond liquid depth, D (m)
Net evaporation, e (mm/day)
First-order rate constant for BOD removal at 20°C, k1 (day–1)
Arrhenius constant, φ
Non-algal fraction of effluent BOD, Fna
Alternatively, in place of P, B and q, we could simply consider Li and Q.
Each parameter is assigned a range of design values, rather than a single
fixed (ie ‘certain’) value, although any parameter could be assigned a fixed
value, for example, the depth D could be set at the single value of, say, 1.5 m.
The range of values could be, for example:
1.5 (fixed value)
A multi-trial Monte Carlo design simulation is now established. This
procedure selects at random a value for each design parameter within the
specified range and then determines the design – that is, it calculates, for
example, the pond mid-depth area and the pond effluent filtered BOD. It then
repeats this ‘random value design procedure’ any required number of times;
usually a 1000-trial design is used. Clearly this cannot be done manually, so
124 Domestic Wastewater Treatment in Developing Countries
an Excel-based design program is used (Sleigh and Mara, 2003a*). The
program completes the 1000 designs and the output can be analysed to provide
frequency histograms and cumulative frequency curves for the pond area,
retention time and the pond effluent filtered BOD. The pond area and
retention time which correspond to the achievement of the required effluent
filtered BOD on either a 50-percentile or a 95-percentile basis (depending on
the regulator’s requirement) are selected as the design values, and this area and
retention time will be sufficient for the parameter ranges chosen to produce an
effluent filtered BOD of the required value for 50 or 95 per cent of the time.
The procedure is illustrated in the second Design Example at the end of this
Microbiological quality
At this stage in the design of a WSP system (ie after the design of the anaerobic
and secondary facultative ponds, or of the primary facultative pond), it is often
worthwhile to assess effluent quality in terms of suitability for re-use in
‘restricted’ irrigation (Chapter 4) – that is for numbers of human intestinal
nematode eggs and E coli.
Human intestinal nematode egg removal
The World Health Organization (1989*, 2004*) recommends that treated
wastewaters used for crop irrigation should contain ≤1 human intestinal
nematode egg/l or, when children under 15 years are exposed, ≤0.1 egg/l
(Chapters 4 and 21).
Using data from WSP in Brazil, India and Kenya, Ayres et al (1992*) found
that there was no statistically significant difference between egg removal in
anaerobic, facultative and maturation ponds (this is to be expected as their
removal mechanism is sedimentation – Chapter 12) (see also von Sperling et
al, 2002). They derived the following equation for percentage egg removal:
R = 100[1 – 0.14exp(– 0.38θ)]
where R is the percentage egg removal in anaerobic, facultative or maturation
ponds; and θ is the retention time in the pond, days. The equation
corresponding to the lower 95 per cent confidence interval for the data used to
derive equation 11.11 is:
R = 100[1 – 0.41exp(– 0.49θ + 0.0085θ2)]
Equation 11.12 is recommended for use in design. It is applied sequentially to
each pond in the series – at this stage in the design first to the anaerobic pond
and then to the secondary facultative pond. Its use is illustrated in the first
Design example at the end of this chapter.
Facultative Ponds 125
E coli removal
If the number of human intestinal nematode eggs in the effluent from the
facultative pond is determined from equation 11.12 as ≤1 per litre, or ≤0.1 per
litre if children under the age of 15 are exposed, then the number of E coli in
the facultative pond effluent should be determined as described in Chapter 12.
For restricted irrigation E coli numbers should be ≤105 per 100 ml (World
Health Organization, 2004*).
The algal biomass in healthy facultative ponds is ~500–2000 mg chlorophyll a
per litre, and thus healthy (ie not overloaded) facultative ponds are dark green
in colour. The concentration of chlorophyll a (C55H72N4O5Mg) is used as algae
vary considerably in volume, by up to three orders of magnitude, so that algal
cell counts are not very informative. As all WSP algae are composed of ~1.5
per cent chlorophyll a by weight, expression of algal biomass in terms of
chlorophyll a, their principal photosynthetic pigment, is more meaningful and,
as chlorophyll a is easily measured, more reproducible (Pearson, 1987*, details
a method for determining chlorophyll a concentrations based on a membrane
filtration, chlorophyll extraction in 90 per cent methanol and measurement of
absorbance at 663 nm).
Chlorophyll a absorbs photons of violet–indigo–blue and orange–red light
of wavelengths ~350–450 and ~650–700 nm (absorption of light of wavelengths 450–650 nm – that is mainly green light, is very weak – hence the green
colour of algae and most higher plants). This absorbed light energy is used to
fix carbon dioxide and to produce oxygen (as a by-product of photosynthesis)
from water. A more complete equation for algal photosynthesis than the simple
equation given in Chapter 9 is (Oswald, 1988b):
106CO2 + 236H2O + 16NH4 + HPO 4 C106H181O45N16P + 118O2 +
171H2O + 14H+
This equation shows that 1 g of algae produces (3776/2429) = 1.55 g of
oxygen. This quantity of oxygen is sufficient to satisfy the oxygen demand of
1.55 g of BODu, which approximately equals 1 g of BOD5 (Chapter 1).
Algal productivity
In-pond algal productivity can be estimated by the ‘light-and-dark-bottle’
technique (American Public Health Association, 1998 – method 10200 J.2):
samples are taken from the pond at 15-cm depth intervals and each is used to
completely fill three BOD bottles – one is used to measure the dissolved oxygen
(DO) concentration immediately (ie at zero time); one of the other two is
covered in aluminium foil (the ‘dark bottle’), and this and the remaining bottle
(the ‘light bottle’) are resuspended in the pond at the depth from which the
126 Domestic Wastewater Treatment in Developing Countries
samples they contain were collected. After a certain time t they are removed
from the pond and the DO concentration in their contents is measured (the
time t that the bottles are immersed in the pond has to be chosen carefully,
usually by trial and error, so that the DO concentration in the dark bottle is
not zero after t minutes and the DO concentration in the light bottle is not
higher than the maximum DO meter reading – usually 20 mg/l; t also depends
on the time of day: longer times are needed early in the morning, or late in the
afternoon, than at midday). The increase in DO in the light bottle (which is
due to algal photosynthesis less bacterial respiration) plus the decrease in the
dark bottle (due to algal and bacterial respiration) is the gross algal DO
production at that level in the pond and at that time of day. The gross DO
production in the pond water column at that time of day is the sum of the
values obtained at each depth; if this is denoted by Σ(GDOP) in mg O2 per
litre, then on an areal basis in kg O2/ha day the gross algal oxygen production
(GAOP) is given by:
108 24 x 60
where V is the volume of the sample bottle, ml; A is the cross-sectional area of
the sample bottle, cm2; and t is the time of immersion of the bottles in the
pond, min. The factor 106 converts mg/l to kg/l; 1000 converts ml to l; 108
converts cm2 to ha; and (24 x 60) converts minutes to days – so that GAOP is
expressed in kg/ha day.
The test needs to be repeated at different times of the day to obtain an
estimate of total daily gross algal oxygen production. For a facultative pond in
northeast Brazil loaded at 250 kg BOD/ha day at 25°C, this was just over 400
kg O2/ha day; and in a maturation pond loaded at 50 kg BOD ha day it was
nearly 300 kg O2/ha day.
The light-and-dark-bottle test described above can be rather ‘tricky’ and
experience is needed to obtain good results.
Algal biomass concentration
Algal biomass concentrations in facultative ponds, and to a lesser extent in
maturation ponds, are influenced by several factors, including light intensity,
the BOD loading on the pond (λs), the un-ionized ammonia (ie NH3)
concentration and the un-ionized H2S concentration. If there are industrial
effluents present in the wastewater, the algae may be adversely affected, even
killed, by industrial toxicants, although many of these are degraded or
removed in anaerobic ponds (Chapter 10).
Light intensity
Light intensity can be expressed in several ways, but in relation to
photosynthesis it is best to express irradiance (radiant solar energy flux density,
Facultative Ponds 127
W/m2) in terms of photon flux density – that is, the number of photons (quanta
of light) falling on a surface of unit area per unit of time. The ‘photosynthetic
photon flux density’ (PPFD) is the number of photons of photosynthetically
active radiation (PAR) – that is, light in the 400–700 nm wavelength range –
falling on 1 m2 of the earth’s surface per second. The units of PPFD are
einsteins/m2 second (but usually µE/m2 s), where one einstein is 6.023 x 1023
photons (6.023 x 1023 is Avogadro’s number). The energy (E, joules) of one
einstein depends on the wavelength of the photons which comprise it:
E = 6.023 x 1023(hc/λ)
where h is Planck’s constant (6.626 x 10–34 J s); c is the speed of light (3 x 108
m/s); and λ is the wavelength of the light, m. For example, for light of
wavelength 663 nm , (ie 6.63 x 10–7 m) the energy of one einstein is given by:
E = (6.023 x 1023)(6.626 x 10–34)(3 x 108)/(6.63 x 10–7) = 0.18 MJ
Once the photons enter water (eg a facultative or maturation pond) their speed
decreases by a factor of 1.33 (the refractive index of water), so the energy of
one einstein of red light at 663 nm is 0.14 MJ. However, the light available for
photosynthesis in water bodies (including WSP) covers the whole PAR
spectrum of 400–700 nm, and the energy of one einstein of underwater PAR is
~0.24 MJ (Kirk, 1994).
The chemical free energy needed to convert one mole (ie gram molecule,
which is the molecular weight in daltons expressed in grams) of CO2 to
carbohydrate is given by:
4 x 1.25 x 0.0964 = 0.48 MJ
where 4 is the number of electrons involved in the reaction; 1.25 is the redox
potential difference between CO2 fixation and O2 evolution, V; and 0.0964
converts energy expressed in eV (electron volts) to energy in megajoules (MJ).
Two photons of light energy are needed per electron, so the light energy
needed by the WSP algae to fix one mole of CO2 is 8 einsteins, which is (8 x
0.24) = 1.92 MJ. Thus the efficiency of algal photosynthesis is
100(0.48/1.92) = 25 per cent. This efficiency is only for CO2 fixation, but
the algae have to use additional light energy to produce organic compounds
such as proteins, lipids and nucleic acids, and so need more than 8 einsteins
of light energy per mole of CO2 fixed, and thus their efficiency is well below
25 per cent. In the ‘real world’ of facultative ponds, characterized by high
turbidity (which is mostly due to the algae themselves), this efficiency can be
as low as 10 per cent – that is, a requirement of 20 einsteins of underwater
PAR per mole of CO2 fixed. Since 1 mole of CO2 fixed is accompanied by
the production of 1 mole (ie 32 g) of O 2, 20 einsteins of light produces
enough oxygen to remove 32 g of BODu equivalent to 21 g of BOD5. Thus a
128 Domestic Wastewater Treatment in Developing Countries
BOD 5 removal of 1 kg/ha day requires 950 (ie ~1000) einsteins of
underwater PAR/ha day.
Meteorological stations generally measure solar radiation intensity in
W/m2 (using an instrument called a pyranometer). One W/m2 = 1 J/m2 s =
864 MJ/ha day. Only ~40 per cent of total solar radiation incident on a calm
water surface becomes underwater PAR, so 1 W/m2 becomes ~346 MJ of
underwater PAR/ha day or, as 1 einstein of underwater PAR = 0.24 MJ, ~1440
einsteins of underwater PAR/ha day. Since ~1000 einsteins of underwater PAR
are required per kg BOD5 removed, the removal of 1 kg BOD5/ha day requires
~0.7 W/m2. Thus, in the warm tropics (25°C), a design BOD5 loading of 350
kg/ha day, with an expected non-algal BOD5 removal of 90 per cent, requires
~220 W/m2 – a level of total solar radiation always exceeded in these regions.
In temperate climates (for example, Europe or New Zealand, with a winter
design temperature of ≤8°C – Abis and Mara, 2003*), a loading of 80 kg/ha
day with a similar removal requires ~50 W/m2, which is almost always
exceeded, even in winter. (Some meteorological stations may still measure solar
radiation intensity in the old units of langleys/day. One langley/day = 1
kcal/cm2 day = 4.168 J/cm2 day = 0.48 W/m2.)
The relationship between light intensity and algal photosynthesis is similar
to that between substrate concentration and bacterial growth rate (equation
3.4). Light at too low an intensity limits photosynthesis, and light at too high
an intensity also limits photosynthesis (‘photoinhibition’). In tropical and
subtropical facultative and maturation ponds the algae are ‘light-saturated’ (ie
light intensity does not limit photosynthesis) at ~50 W/m2, and photoinhibition
sets in at ~300 W/m2 (this level of photoinhibition explains why the algal band
in stratified facultative ponds moves up and down within the top 50 cm of the
pond as the algae seek to remove themselves from inhibitory light intensities).
BOD loading
The BOD loading applied to facultative ponds has a major influence on the inpond algal biomass. In primary facultative ponds in northeast Brazil at 25°C
chlorophyll a levels fell sharply with increasing BOD load up to ~450 kg/ha
day, above which they were essentially constant at ~200 µg/l (Figure 11.4).
This was the reason for selecting a design BOD loading rate of 350 kg/ha day
at 25°C (equation 11.3). Along with higher BOD loads, higher BOD loading
rates also bring higher loadings of sulphate and ammonia.
Sulphide toxicity
Sulphates in raw wastewater are reduced in anaerobic ponds and primary
facultative ponds by the obligately anaerobic sulphate-reducing bacteria to
sulphides, and these are also present in secondary facultative ponds which
receive them in the anaerobic pond effluent. Un-ionized H2S is highly toxic to
pond algae and, as shown in Figure 10.2, the amount of total sulphide present
as dissolved H2S gas increases as the pH decreases. Dissolved H2S rapidly
passes through the algal cell membrane to attack its photosynthetic apparatus;
hence algal growth and algal oxygen production are inhibited, and the pond
Facultative Ponds 129
BOD surface loading (λs, kg/ha day)
BOD loading rate (kg/ha day)
Source: Silva (1982)
Figure 11.4 Variation of Chlorophyll a with Surface BOD Loading on
Primary Facultative Ponds in Northeast Brazil
quickly turns anaerobic or at least anoxic. This toxicity increases with
decreasing pH. Experiments with strains of four common facultative pond
algae – Chlamadomonas, Chlorella, Euglena and Scenedesmus – showed that
photosynthesis is inhibited by 50 per cent (relative to that in the absence of
H2S) by very low levels of H2S which vary from alga to alga: Chlamadomonas
(45 µM), Chlorella (30 µM), Scenedesmus (22 µM) and Euglena (10 µM)
(1 µM H2S = 34 µg H2S/l) (Pearson et al, 1987b). The comparatively high
tolerance of Chlamadomonas to H2S most probably explains why it is
sometimes found as a surface film on anaerobic ponds (Chapter 10).
In primary facultative ponds in northeast Brazil loaded at 160–580 kg
BOD/ha day at 25°C, Silva (1982) found a linear relationship between
chlorophyll a and sulphide:
[Chl.a, µg/l] = 2550 – 346[S, mg/l]
This low tolerance of facultative pond algae to sulphide indicates that
anaerobic pond effluent should not be introduced at or near the surface of a
secondary facultative pond, since this is where the algae are. Instead it should
be introduced in the lower part of the pond (at least 0.8 m below the surface)
where the H2S it contains (5–15 mg/l) can be oxidized by anaerobic
photosynthetic bacteria (see below under ‘Purple ponds’).
130 Domestic Wastewater Treatment in Developing Countries
Ammonia toxicity
Ammonia is present in aqueous solution as dissolved ammonia gas (NH3 – ie
un-ionized ammonia) and the ammonium ion (NH4). As with sulphide, it is
the un-ionized ammonia that is toxic to algae, but with one important
difference: as the pH increases, the proportion of total ammonia present as
NH3 also increases, thus ammonia toxicity increases with increasing pH.
Pearson et al (1987b) found that 50 per cent inhibition of photosynthesis in
the four pond algae occurred in a different order from that caused by H2S:
Chlorella (3.9 mM), Scenedesmus (1.6 mM), Euglena and Chlamadomonas
(both 0.9 mM) (1 mM NH3 = 14 mg NH3–N/l).
As ammonia toxicity increases with pH, it can be self-correcting at
sublethal concentrations. This is because rapid photosynthesis leads to a high
pond pH (often 9–10), ammonia toxicity thus increases and therefore
photosynthesis decreases; the pH falls and the ammonia toxicity decreases –
this leads to increased photosynthesis, an increase in pH and therefore in
ammonia toxicity, and so the cycle repeats itself.
When facultative ponds are overloaded, either permanently or transiently, they
often appear purple in colour (although the precise colour can vary from pink,
through red and purple, to light brown). This is due to the toxic effects of
mainly sulphide on the algae and the consequent predominance of purple and
green anaerobic photosynthetic bacteria. These bacteria use light energy to fix
carbon dioxide with the concomitant oxidation of hydrogen sulphide to
sulphur and sulphate:
21H2S + 10CO2 + 2NH3 2(C5H8O2N) + 21S + 16H2O
21S + 30CO2 + 6NH3 + 36H2O
6(C5H8O2N) + 21H2SO4
These equations are for ‘photo-autotrophic’ growth (ie light-enabled growth
on carbon dioxide). The green and purple photosynthetic bacteria can also
utilize simple organic compounds, such as acetate, for ‘photoheterotrophic’
H2S + 5CH3OOH + 2NH3
2(C5H8O2N) + S + 6H2O
S + 15CH3COOH + 6NH3
6(C5H8O2N) + H2SO4 + 14H2O
Anaerobic bacterial photosynthesis does not result in the production of oxygen
(ie it is ‘anoxygenic’) as H2S is oxidized, rather than H2O as in the case of algal
There are many photosynthetic bacteria, but two main groups are
important in facultative ponds: the purple sulphur bacteria belonging to the
Facultative Ponds 131
Figure 11.5 Photosynthetic Purple Sulphur Bacteria: top, Thiopedia sp;
bottom, Chromatium sp
132 Domestic Wastewater Treatment in Developing Countries
family Chromatiaceae, and the green sulphur bacteria belonging to the family
Chlorobiaceae. The Chromatiaceae deposit sulphur granules inside their cells
(Figure 11.5), and the Chlorobiaceae deposit them outside their cells (the
pathways in the second equations in the above two pairs of photosynthesis
equations are only used when hydrogen sulphide is unavailable; in times of
H2S plenty, S is stored for use in times of H2S famine).
Bacteriochlorophylls (the principal pigments of photosynthetic bacteria)
absorb light of a longer wavelength than that absorbed by algal chlorophyll
(750–900 nm – ie in the infra-red part of the spectrum, rather than <700 nm).
Thus, in facultative ponds that are not overloaded, the anaerobic
photosynthetic bacteria are found below the algae (longer-wavelength light
penetrates deeper than shorter-wavelength light) and closer to the anaerobic
sulphate-reducing bacteria that produce the sulphide – that is, they act as a
sulphide ‘filter’, protecting the algae from the toxic effects of sulphide and also
reducing odour release. They are thus very important in both correctly loaded
and overloaded facultative ponds.
Gently mixing (stirring, circulating) the contents of an overloaded primary
facultative pond can greatly aid its performance – often to the point where it
no longer acts as if it were overloaded. The use of wind-powered mixers
(sometimes called ‘wind-powered aerators’) can be a cost-effective means to
mix these ponds; alternatively, low-powered electric mixers can be used (the
power input is <1 W/m3, rather than the 3–6 W/m3 used in aerated lagoons –
Chapter 20). However, a generally more appropriate strategy in warm climates
is to install an anaerobic pond ahead of the facultative pond, although, of
course, wind- or electric-powered mixers could be used on overloaded
secondary facultative ponds.
Viraraghavan et al (2002) studied the performance of a 5.2-ha, 1.5-m deep
facultative pond, serving a village of 565 people in central Canada, both with
and without two ‘Little River Pond Mills’ wind-powered mixers (Sunset Solar
Systems, 2002*). The influent BOD of 170 mg/l was reduced to 40–80 mg/l
without the mixers and to 10–20 mg/l with them. The estimated oxygen
transfer rate was equivalent to ~1.5 kg O2/kWh at a wind speed of ~5 m/s.
Further discussion on the use of wind/electric-powered mixers for primary
facultative ponds is given by Mara (2003*).
Conventional design
Design (a) a primary facultative pond to treat the wastewater detailed in the
Design example in Chapter 10; and (b) a secondary facultative pond to treat
the effluent from the anaerobic pond designed in Chapter 10.
Facultative Ponds 133
Additional design parameters: e = 5 mm/d; Ei = 500 human intestinal
nematode eggs/l; and Ni = 5 x 107 E coli per 100 ml.
Primary facultative pond
From equation 11.3 λs = 350 kg/ha day for 25°C. Thus from equations 11.1
and 11.7 and assuming a depth of 1.5 m:
Af = 10LiQ/λs
= (10 x 300 x 10,000)/350
= 85,720 m2 (ie 8.6 ha)
θf = 2AfDf/(2Q – 0.001eAf)
= (2 x 85,720 x 1.5)/[(2 x 10,000) – (0.001 x 5 x 85 720)]
= 13 days
Secondary facultative pond
With Li = 90 mg/l (the BOD of the anaerobic pond effluent), equations 11.1
and 11.7 give:
Af = 25,720 m2
θf = 3.88 days
The retention time is <4 days, the minimum retention time in facultative ponds.
Therefore recalculate Af from equation 11.7 with θf = 4 days:
Af = 2Qθf/(2D + 0.001eθf)
= (2 x 10,000 x 4)/[(2 x 1.5) + (0.001 x 5 x 4)]
= 26,500 m2 (ie 2.7 ha)
Check for BOD and helminth egg removal in the anaerobic and secondary
facultative ponds:
BOD removal: from equations 5.7 and 5.8, the BOD of the effluent from the
secondary facultative pond is:
k1 = 0.1 (1.05)T–20
= 0.1 (1.05)25–20
134 Domestic Wastewater Treatment in Developing Countries
= 0.13 days–1
Le = Li/(1 + k1θf)
= 90/[1 + (0.13 x 4)]
= 60 mg/l (unfiltered BOD)
Thus the filtered BOD is given by equation 11.8 as:
Le (filtered) = 0.3 (Le, unfiltered)
= 0.3 x 60
= 18 mg/l
Thus the effluent from the secondary facultative pond is suitable for surface
water discharge.
Egg removal: using equation 11.12, the egg removal in the 1-day anaerobic
pond is 75 per cent, and in the 4-day secondary facultative pond 93 per cent.
Thus the number of eggs in the effluent from the latter is:
Ee = 500(1 – 0.75)(1 – 0.93)
= 9 per litre
This is >1 per litre, the guideline level for restricted irrigation, and >>0.1 per
litre, the recommended level for restricted irrigation when children under the
age of 15 years are exposed (Chapter 21). Thus this particular secondary
facultative pond effluent is not suitable for restricted irrigation, and it requires
further treatment in maturation ponds (Chapter 12).
Effluent flow: the flow from the secondary facultative pond, which is required
for the design of the maturation ponds, is given by:
Qe = Qi – 0.001eAf
= 10,000 – (0.001 x 5 x 26,500)
= 9868 m3/day
Land saving achieved by anaerobic ponds
The areas calculated above are:
Facultative Ponds 135
anaerobic and secondary facultative ponds: 3340 + 26,500 = 29,840 m2
primary facultative pond: 85,720 m2.
Thus the primary facultative pond requires 187 per cent more land than the
anaerobic and secondary facultative ponds combined. This illustrates very well
the following observation of Marais (1970): ‘Anaerobic pretreatment is so
advantageous that the first consideration in the design of a series of ponds
should always include the possibility of anaerobic pretreatment’.
Design based on uncertainty
This design procedure is based on von Sperling (1996b*). The following ranges
of parameter values were used for the design of the secondary facultative pond
(of course, different ranges could be used and different designs obtained):
8000–10,000 m3/day
90–112 mg/l
1.5 m (fixed value)
4–6 mm/day
0.09–0.11 day–1
The three equations of interest are equations 11.1 and 11.3 combined,
equation 11.7, and equations 5.7, 5.8 and 11.8 combined – that is:
Af = 10LiQ/[350(1.107 – 0.002T)T–25]
θf = 2AfDf/(2Q – 0.001eAf)
Le (filtered) = FnaLi/[1 + (k1(20)φT–20θf)]
The Monte Carlo programme (Sleigh and Mara, 2003a*) produced the
following results after 1000 trials (each 1000-trial run produces slightly
different results):
Mean area = 26,046 m2 and mean filtered Le = 6 mg/l, and
95-percentile area = 30,524 m2 and 95-percentile filtered Le = 7 mg/l.
If the regulator requires compliance with a mean effluent filtered BOD, then
the pond area is 2.6 ha. If, however, compliance were required on a 95percentile basis, the area would be 3.1 ha.
Maturation Ponds
The main function of maturation ponds is to reduce the number of excreted
pathogens, principally faecal bacteria and viruses, present in the effluent of
facultative ponds (Chapter 11) to a level suitable for agricultural and/or
aquacultural re-use (Chapters 21 and 22). BOD and suspended solids are
removed only very slowly, and nutrient (nitrogen and phosphorus) removal is
also quite slow, although in a well-designed and properly operated and
maintained series of WSP (anaerobic, facultative and several maturation
ponds), cumulative removals of BOD, suspended solids and nutrients are high,
as shown in Table 12.1. The extremely high removal of E coli in WSP is also
shown in Table 12.1: from 5 x 107/100 ml in the raw wastewater to 30/100 ml
in the effluent of the third maturation pond – a removal of 6 log units, or
99.9999 per cent (the effluent of the second maturation pond contained <1000
E coli per 100 ml, and so would be suitable for unrestricted irrigation –
Chapters 4 and 21; thus there would be no point in practice in having the third
maturation pond, the effluent of which contained only 30 E coli per 100 ml –
better bacteriologically than the water used for drinking by many millions of
people in developing countries).
Maturation ponds are typically aerobic throughout their depth, and they
show less stratification than facultative ponds. They contain a greater diversity
of algal genera (Table 11.1), but algal biomass is lower. In a series of
maturation ponds, algal diversity increases, but algal biomass decreases, from
pond to pond. Depths are typically 1 m; shallower ponds achieve higher faecal
bacterial and viral removals due to greater light penetration, but unlined ponds
with depths of less than 1 m are likely to contain emergent macrophytes
(rooted plants growing up from the pond base) – these provide a suitably
shaded habitat for mosquito breeding, and so should be avoided by having
depths ≥1 m unless the ponds are lined (Chapter 13).
Maturation Ponds 137
Table 12.1 Performance of a Series of Five WSP in Northeast Brazil
Parameter Raw
Anaerobic Facultative
(mg/l)a wastewater pond
pond Maturation Maturation Maturation percentage
Ammonia-N 45
coliforms5 x 107
3 x 106
3 x 105
2 x 104
Notes: The ponds had retention times of 5.5 d, except the anaerobic pond (6.8 d) and
maturation pond 3 (5.8 d). The results are mean values over a 24-month period.
Except chl.a (µg/l) and FC (per 100 ml)
Total P for raw wastewater and anaerobic pond effluent; soluble (ie, non-algal) P for others
Source: Silva (1982)
The mechanisms underlying viral die-off and removal in WSP are not fully
understood, but it is believed that they are removed mainly by sedimentation
following adsorption on to solids, including algae (which settle out when they
die). Few studies have been done on viral removal in WSP (as counting viruses
is expensive and requires a highly skilled microbiologist). Oragui et al (1995*)
studied rotavirus and faecal coliform removal in several WSP series in
northeast Brazil, each of which comprised an anaerobic pond, a secondary
facultative pond and three maturation ponds, with overall retention times of
10–25 days. Rotaviruses were reduced from 1 x 104/l of raw wastewater to
<2/l of final effluent – a removal of 99.997 per cent; the corresponding faecal
coliform removal was from 3 x 107/100 ml to ~50/100 ml, a removal of
99.9998 per cent. Rotavirus removal was best in the 0.6 m deep tertiary
maturation ponds which had retention times of 5 days: rotavirus numbers
decreased from 1500 per litre to <2/l, a removal of ~99.9 per cent. The data in
Table 12.2 for rotavirus removal indicate first-order rate constant values (for
use in equation 12.1 below) of ~3 day–1 in the anaerobic pond and ~0.3–0.5
day–1 in the facultative and maturation ponds at 25°C.
Bacteriophages, which are viruses that attack bacteria, can be used as
surrogates for human viral removal in WSP. Vorkas and Lloyd (2000*),
working on WSP near Cali, Colombia, showed that phages of Erwinia,
138 Domestic Wastewater Treatment in Developing Countries
Pseudomonas and Serratia were good models for viral transport through, and
removal in, WSP. Phage die-off was higher at in-pond pH values >8.5 and was
accelerated by the same sunlight-mediated effects that Curtis et al (1992a)
found for faecal bacterial removal in WSP (see below).
The faecal bacteria of interest are the pathogens which cause infectious
intestinal disease, such as Campylobacter, diarrhoeagenic Escherichia coli,
Salmonella, Shigella and Vibrio cholerae (Chapter 2), and E coli which is used
as a bacterial and viral pathogen indicator organism (Chapter 3). The removal
of faecal bacteria in WSP has been studied by many research workers, and it is
almost always found that their removal is very high in a well designed,
properly operated and maintained series of WSP. As an example of this huge
research effort over the past 30 or so years, the removal data of Oragui et al
(1987) for a 5-pond series in northeast Brazil are given in Table 12.2. This
data set is unique in that it contains data from the same series of WSP and at
the same time on the removal of faecal coliforms, two bacterial pathogens and
two viral pathogens. It is also highly supportive of the World Health
Organization’s (WHO, 1989*, 2004*) guidelines for unrestricted irrigation
(Chapters 4 and 21) as it shows that, when the faecal coliform count was
reduced to 7000/100 ml, both bacterial pathogens had been completely
removed and that only very low numbers of the viral pathogens were present –
thus demonstrating that the WHO guideline level of ≤1000 E coli per 100 ml
is perfectly adequate (this point is discussed in greater detail in Chapter 21).
Oragui et al (1993) were able to monitor the removal of V cholerae O1 in
WSP when the seventh cholera pandemic arrived in northeast Brazil. The
(slightly unusual) WSP series comprised a 1-day anaerobic pond followed by
nine 2-day ponds. The raw wastewater contained 485 V cholerae per litre and
2 x 107 faecal coliforms per 100 ml. Most (94 per cent) of the V cholerae were
removed in the anaerobic pond (Chapter 10), and complete (ie 100 per cent)
removal was achieved in the effluent of the fifth 2-day pond (ie after 11 days);
the corresponding faecal coliform count was 6 x 104/100 ml – again supporting
the WHO E coli guideline for unrestricted irrigation.
Several factors have been proposed to explain the die-off of faecal bacteria
in WSP, and they can be conveniently grouped into ‘dark-mediated’ processes
and ‘light-mediated’ processes. The dark-mediated processes (which are really
‘light-independent’ processes, as they occur in both the light and the dark),
include sedimentation of the faecal bacteria adsorbed on to settleable solids or
contained within flocs of settleable solids; predation (ie consumption) by freeliving protozoa and micro-invertebrates such as Daphnia and Moinia (water
fleas); and death due to starvation and senescence. There are external factors
which aid these dark-mediated factors, for example, pond depth (the deeper a
pond, the greater the proportion of it that is in darkness), and organic loading
(the greater the loading, the more anaerobic the pond and the effects of the
light-mediated factors are reduced).
Maturation Ponds 139
Table 12.2 Bacterial and Viral Removals in a Series of Five WSP in
Northeast Brazil
Anaerobic Facultative
wastewater pond
pond maturation maturation maturation
Faecal coliforms
2 x 107
1 x 104
4 x 106
6 x 103
8 x 105
1 x 103
2 x 105
3 x 104
7 x 103
Notes: The anaerobic pond had a retention time of 1 day, each of the others 5 days.
a Bacterial numbers per 100 ml, viral numbers per 10 l.
Source: Oragui et al (1987)
The most important of the light-mediated factors are:
Time and temperature: these are taken as light-mediated factors as the
pond temperature is the result of the sunlight intensity at the pond surface,
and the longer the time that the pond is exposed to the light-mediated
factors, the greater the faecal bacterial die-off. (There is potential for
confusion here, as we know that bacterial growth rates increase with
temperature, but so do bacterial death rates. Whatever the cells are going
to do – grow or die – they do it more quickly at higher temperatures.) Time
and temperature are the two factors included in Marais’ (1974) model for
E coli die-off in WSP (see equations 12.1 and 12.2 below).
High pH: in-pond pH values ≥9.4 induce very rapid faecal bacterial dieoff (Parhad and Rao, 1974; Pearson et al, 1987c), except that of V cholerae
(for which pH 9 is used in isolation media to inhibit other faecal bacteria;
as noted in Chapter 10, V cholerae is very sensitive to sulphide
concentrations ≥3 mg/l which occur in anaerobic ponds). High pH is a
light-mediated factor as it is induced by the pond algae. When algae are
photosynthesizing rapidly, their demand for CO2 outstrips its supply from
bacterial metabolism and transfer across the pond surface; this absence of
dissolved CO2 in the pond disturbs the CO2–bicarbonate–carbonate
equilibrium, and consequently bicarbonate and carbonate ions dissociate,
as follows:
2HCO3 CO 3 + H2O + CO2
CO 3 + H2O 2OH + CO2
The resulting CO2 is fixed by the algae, and the hydroxyl ions (OH )
accumulate to raise the pH. High pH values ≥9.4, even >10, occur on
sunny days close to the pond surface, which is therefore where the most
140 Domestic Wastewater Treatment in Developing Countries
rapid faecal bacterial die-off occurs. A high in-pond (ie extracellular) pH
kills faecal bacteria by making them unable to maintain their optimal
intracellular pH of 7.4–7.7.
High light intensity and high dissolved oxygen: the role of light in
mediating faecal bacterial die-off was elucidated by Curtis (1990) (see
also Curtis et al, 1992a, 1992b, 1994; see also Davies-Colley et al,
2000*). Curtis found that light of wavelengths up to 700 nm could
damage faecal bacteria; however, light of wavelengths below 425 nm (ie
including UV light) was unimportant in WSP as it is almost wholly
absorbed in the first few mm of the pond. Light of wavelengths >425 nm
could only damage faecal bacteria in the presence of both a dissolved
sensitizer such as the humic substance gilvin (‘dissolved yellow matter’)
and dissolved oxygen – both gilvin and oxygen are required for lightinduced damage of faecal bacteria (‘photo-oxidation’). Gilvin is present
in almost all waters, including wastewaters and WSP (it is measured by
the absorbance of a 0.2-µm filtrate of the water sample at 440 nm). The
light–oxygen–gilvin damage is enhanced by intracellular pH values >7.7,
so the pond algae are crucial for the die-off of faecal bacteria in WSP:
they produce high dissolved oxygen levels and induce high in-pond pH
values which induce an intracellular pH >7.7, which in turn and in
conjunction with high light intensities (>~500 W/m2) achieves rapid faecal
bacterial die-off. The way in which the combination of high light intensity,
high dissolved oxygen, high pH and gilvin kills faecal bacteria appears to
be as follows: gilvin absorbs the light and then reacts with oxygen to form
oxygen radicals (eg hydrogen peroxide) which damage the cell membrane
and so cause the cell to die; and the high pH enhances cell damage in the
way explained above.
Helminth eggs and protozoan cysts
Eggs and cysts are removed in WSP by sedimentation, and thus their removal
occurs mainly in the first ponds in a series – anaerobic and facultative ponds.
Settling velocities for eggs and cysts are given in Table 12.3. Cysts are smaller
and settle more slowly than eggs, and hence their removal requires a longer
retention time. Grimason et al (1993) studied the removal of Cryptosporidium
Table 12.3 Settling Velocities for Parasite Eggs and Cysts
Egg/cyst size (µm)
55 x 40
22 x 50
60 x 40
Relative density
Settling velocity (m/h)
Sources: Shuval et al (1986*) and Department of the Environment and Department of Health
Maturation Ponds 141
Table 12.4 Helminth Egg Removal in Waste Stabilization Ponds in
Northeast Brazil
Anaerobic Facultative
wastewater pond
pond maturation maturation maturation
Egg numbers
per litre
Retention time,
Egg numbers
per litre
Retention time,
Source: Silva (1982)
oocysts and Giardia cysts in 11 WSP systems in Kenya with overall retention
times of 14–133 days. The raw wastewaters contained 13–73 oocysts and
213–6213 cysts/l, but none was found in the final effluents, except at one
location – the overloaded WSP at Eldoret (actual retention time 22 days; design
value 33 days), the effluent of which had 40–50 Giardia cysts/l. Amahmid et
al (2002*) found complete removal of Giardia cysts in the WSP at Marrakech,
Morocco which had a retention time of 16 days.
Helminth eggs are completely removed in WSP within around 5–15 days
(Table 12.4), depending on the number of eggs in the raw wastewater.
Equation 11.11 of Ayres et al (1992*), based on egg removal in WSP in Brazil,
India and Kenya, describes egg removal very well (this was independently
confirmed by von Sperling et al, 2002b), but for design purposes equation
11.12 should be used.
Marais’ method
Marais (1974) refined the earlier model of Marais and Shaw (1961) for E coli
removal in WSP which was modelled on first-order kinetics in a completely
mixed reactor (Chapter 5). The resulting equations for a single pond are the
following modifications of equations 5.7 and 5.8:
Ne =
1 + kB(T)θ
142 Domestic Wastewater Treatment in Developing Countries
kB(T) = 2.6(1.19)T–20
where Ne and Ni are the numbers of E coli per 100 ml of the pond effluent
and influent, respectively; kB(T) is the first-order rate constant for E coli
removal at T°C in a completely mixed reactor, day–1; θ is the mean hydraulic
retention time in the pond, days; and T is the design temperature, °C.
The design temperature is the mean temperature of the coolest month in
the season in which the maturation ponds have to produce an effluent of the
required microbiological quality. This is the irrigation season if the effluent is
to be re-used in agriculture. The design temperature is the mean temperature
of the coolest month if the effluent is used for fishpond fertilization or if it is
discharged into bathing waters or near shellfisheries (as these are year-round
Equation 12.2 gives the value of kB(T) as 2.6 day–1 at 20°C, and its value is
highly temperature-dependent: it changes by 19 per cent for each change in
temperature of 1 degC.
For a series of WSP comprising an anaerobic pond, a secondary facultative
pond and n equally sized maturation ponds, equation 12.1 is rewritten as:
Ne =
(1 + kB(T)θa) (1 + kB(T)θf) (1 + kB(T)θm)n
where Ne and Ni are now the E coli numbers per 100 ml of the final effluent
and the raw wastewater, respectively; the subscripts a, f and m refer to the
anaerobic, facultative and maturation ponds; and n is the number of equally
sized maturation ponds.
In this chapter it is assumed that it is possible to have n equally sized
maturation ponds. However, if during the physical design stage (Chapter 13),
site conditions are such that this is not possible, then a redesign is necessary
with the term (1 + kB(T)θm)n in equation 12.3 replaced by:
(1 + kB(T)θm1)(1 + kB(T)θm2)…(1 + kB(T)θmn)
The value of θm is subject to the following three constraints:
θm >/ θf,
θm </ θm , and
λs(m1) >/ 0.75λs(f)
The first constraint simply states that the retention time in the maturation
pond should not be greater than that in the facultative pond. There is no
theoretical reason for this, but it seems sensible (ie it is based on ‘engineering
judgement’). The second constraint sets a minimum value for the retention
Maturation Ponds 143
time in the maturation ponds to permit algal reproduction and to minimize
hydraulic short-circuiting. Marais (1974) recommends a value of 3 days for
θm . The third constraint, which is the critical one to consider first, sets a limit
on the BOD surface loading on the first maturation pond; clearly this should
not be more than that on the facultative ponds, and it is better if it is less; here
it is set at 75 per cent of the latter.
Equation 11.1 for surface loading can be rewritten with Q/A = D/θ for the
first maturation pond as:
λs(m1) = 10LiDm1/θm1
Here, Li is the unfiltered BOD of the influent into the first maturation pond –
that is, that of the effluent from the facultative pond (Le(fac)) as determined by
equations 5.7 and 5.8 (with k1(20) = 0.1 day–1 for secondary facultative ponds,
or 0.3 day–1 for primary facultative ponds, and φ = 1.05). Thus equation 12.4
θm1 = 10Le(fac)Dm1/0.75λs(fac)
and equation 12.3 is rewritten as:
Ne = Ni/(1 + kB(T)θa)(1 + kB(T)θf)(1 + kB(T)θm1)(1 + kB(T)θm)n
where θm is now the retention time in the second and subsequent maturation
ponds of which there are now n.
At this stage in the design all the terms in equation 12.6 are known or
assumed, except θm and n: Ne is the effluent quality required (eg 1000
E coli/100 ml for unrestricted irrigation – Chapter 21); Ni is either known or
taken as 5 x 107 per 100 ml; θa and θf are known from the designs for the
anaerobic and facultative ponds (Chapters 10 and 11); and θm1 is calculated
from equation 12.5 and kB(T) from equation 12.2.
A more convenient form of equation 12.6 is:
θm = {[Ni/Ne(1 + kB(T)θa)(1 + kB(T)θf)(1 + kB(T)θm1]1/n – 1}/kB(T)
Equation 12.7 is solved first for n = 1, then for n = 2, and so on, until the
calculated value of θm is <θm ; assume this happens when n = ñ. Consider only
values of θm which are <θf and >θm , but also θm itself; and choose the
combination of n and θm, including ñ and θm , which has the least overall
retention time and therefore the least land area requirement. As an example
(and only as an example), suppose the following solutions to equation 12.7
were obtained:
144 Domestic Wastewater Treatment in Developing Countries
for n = 1, θm = 256 days;
n = 2, θm = 27 days;
n = 3, θm = 4 days; and
n = 4, θm = 1.9 days (ie ñ = 4).
Assuming θf <27 days, then the two combinations to consider are n = 3 and θm
= 4 days, and ñ = 4 and θm = 3 days. Both give an overall retention time of
12 days; in this case it is better to have three 4-day maturation ponds rather
than four 3-day ones in order to minimize embankment construction.
Values of kB(20) and φ
The values of kB(20) and φ given by Marais in equation 12.2 are 2.6 day–1 and
1.19, respectively. Marais derived these values from his analysis of the faecal
coliform data from the primary facultative and maturation ponds serving
‘Community C’ (in the USA) reported by Slanetz et al (1970). The temperature
range for these ponds was 2–21°C; the BOD loading on the primary facultative
pond was not stated, but was presumably ~50 kg/ha day or less.
However, the value of kB is highly dependent on the BOD loading. Figure
12.1 shows a steep decline in the value of kB in primary facultative ponds at
25°C in northeast Brazil up to a BOD loading of 400 kg/ha day, above which
it was essentially constant at ~2.8 day–1, which is its value in anaerobic ponds
at 25°C (equation 12.8 below). Figures 11.4 and 12.1, which have the same
shape, together provide direct experimental evidence that the value of kB is
controlled by the algal concentration (see Figure 12.2), which in turn is
controlled by the BOD loading.
FC removal rate constant (day–1)
BOD loading rate (kg/ha day)
Source: Silva (1982)
Figure 12.1 Variation of kB with Surface BOD Loading on Primary
Facultative Ponds in Northeast Brazil
Maturation Ponds 145
FC removal rate constant (day–1)
In-pond chlorophyll a concentration (µg/l)
Figure 12.2 Variation of kB with In-pond Chlorophyll a Concentration in
Primary Facultative Ponds in Northeast Brazil
As noted by Catunda (1994) and Mendonça (2000), several research workers
have found kB(20) values <2.6 day–1 and φ values <1.19. Examples are given in
Table 12.5. Catunda also reported kB values at 25°C from WSP in the same
city in northeast Brazil that varied significantly with the pond depth: from 7.3
day–1 in 0.3 m deep ponds to 0.8 day–1 in 3 m deep ponds. Thus the original
method of Marais with kB(20) = 2.6 day–1 and φ = 1.19 may not be generally
applicable. A more generally applicable method is that of von Sperling, which
is now discussed.
von Sperling’s method
As noted in Chapter 5, von Sperling (1999*, 2002*, 2003*) recommends the
use of the dispersed flow model (equation 5.13) for E coli removal in WSP,
with equation 5.14 for δ and equations 5.8 and 5.15 for kB. For convenience
these equations are repeated here:
The ‘abbreviated’ Wehner–Wilhelm equation for dispersed flow rewritten
in terms of E coli numbers:
Ne = Ni[4a/(1 + a)2]exp[(1 – a)/2δ]
where a = Ⱦ(1 + 4kB(T)θδ) and θ = V/Q (ie the nominal mean hydraulic
retention time), days.
von Sperling’s equation for δ:
146 Domestic Wastewater Treatment in Developing Countries
Table 12.5 Reported Values of kB(20) and φ for Use in Equation 12.2
kB(20) (day–1)
Klock (1971)
Marais (1974)
Skerry and Parker (1979)
Arceivala (1981)
Mills et al (1992)
Yanez (1993)
Mayo (1995*)
Mara et al (2001b*)a
a For temperatures >20°C
δ = (L/B)–1
von Sperling’s equations for kB:
kB(20) = 0.92D–0.88θ–0.33
kB(T) = kB(20)(1.07)T–20
where L, B and D are the pond length, breadth and depth, respectively, m.
Equation 5.15 was derived by von Sperling from the faecal coliform removal
data from 33 facultative and maturation systems in tropical and subtropical
Brazil (latitude 7–24º South). It thus represents a very wide range of conditions
and can be recommended in preference to Marais’ method, particularly since
more recent workers have, as noted above, found complete-mix values of kB(20)
<2.6 day–1 and φ <1.19.
This design procedure must include the three constraints for the retention
times in the maturation ponds and the judicious selection of the combinations
of retention time in, and number of, maturation ponds (as described above
under ‘Marais’ method’). As recommended by von Sperling (2003*), the whole
procedure can also be subjected to ‘uncertainty’ analysis as described in
Chapter 11 for facultative ponds (see also Gawasiri, 2003).
However, Pearson et al (1995*, 1996b*) found that the performance of
secondary facultative and secondary maturation ponds in northeast Brazil at
25°C was not influenced by either depth (1–2 m for facultative ponds, 0.4–0.9
m for maturation ponds) or length-to-breadth ratio (1–6.5 to 1 for facultative
pond, 2.8–8 to 1 for maturation ponds) (Table 12.6). The actual faecal
coliforms removals in these ponds were very close to those predicted by
Marais’ equations. These ponds in northeast Brazil were optimally loaded, as
Maturation Ponds 147
Table 12.6 Performance Data for WSP with Different Depths and
Length-to-Breadth Ratios in Northeast Brazil at 25 ºC
Length- Depth Retention BOD
In-pond Faecal Unfiltered Ammoniareference to(m)
chl a coliforms BOD
(day) (kg/ha day)(mg/m2)a (per
100 ml)
(a) Secondary facultative ponds
6.5 to 1 1.00
6.5 to 1 1.33
6.5 to 1 1.67
6.5 to 1 2.00
1 to 1
(b) Secondary maturation ponds
8 to 1
2.8 to 1 0.39
2.8 to 1 0.39
2.8 to 1 0.64
2.8 to 1 0.90
11 x 105
9 x 105
9 x 105
8 x 105
9 x 105
a Chlorophyll a units are mg/m2 – ie the algal biomass in the water column below 1 m2 of pond
surface area
b Ponds M18 and M19 were identical (internal experimental controls)
Source: Pearson et al (1995*, 1996b*)
were the ponds in the US that Marais used to derive his equations; in contrast,
many of the pond systems used by von Sperling were not optimally loaded, as
was the case for many of those referred to in Table 12.5. Nevertheless, von
Sperling’s method for E coli removal in tropical and subtropical WSP may be
the most generally applicable model since, in practice, most pond systems are,
or become, overloaded.
For anaerobic ponds, which are often square (ie L/B = 1 and, from
equation 5.14, δ also = 1), the complete-mix equation can be used for E coli
removal (equation 12.1), with kB(T) given by:
kB(T) = 2.0(1.07)T–20
Equation 12.8 was derived from faecal coliform removal data reported by Silva
(1982) for anaerobic ponds in northeast Brazil that had retention times of
0.8–6.8 days.
Light intensity
Xu et al (2002*) also used the dispersed flow model, but they correlated kB(T)
with solar radiation intensity (see Chapter 11) as well as with temperature, as
148 Domestic Wastewater Treatment in Developing Countries
kB(T) = kB(20)(0.915)T–20exp(aIm)
where a is a constant (for values, see below); and Im is the depth-averaged inpond solar radiation intensity, J/cm2 day, given by:
Im = I0 (1 – e–KD)/KD
where I0 is the solar radiation intensity at the pond surface, J/cm2 day; D is the
pond depth, m; and K is the light extinction coefficient, m–1, given by:
K = 24.1 + 0.69(SS)
where SS is the in-pond suspended solids concentration, mg/l. The in-pond SS
in equation 21.11 is basically a surrogate for the in-pond algal concentration
as almost all the SS in maturation ponds are algae.
Equation 12.9 was derived for secondary and tertiary maturation ponds
on the island of Noirmoutier off the French Atlantic coast (temperature,
6–22°C; solar radiation intensity, 300–2400 J/cm2 day – ie 35–280 W/m2);
kB(20) was found to be 0.019 day–1 and a 0.170 for the 1.4-m deep secondary
pond, and 0.065 day–1 and 0.191 for the 2.2-m deep tertiary pond. It shows
that, at least in this location, solar radiation intensity is more important than
temperature in E coli removal in maturation ponds. Research is clearly needed
to develop a form of equation 12.9 for tropical and subtropical maturation
If the facultative pond effluent contains >1 egg/l or, if children under the age of
15 years are exposed (see Chapter 21), >0.1 egg/l, then one or two maturation
ponds are required to reduce the egg count to one of these levels in order that
the effluent can be used for restricted irrigation. The retention time in the first
of these maturation ponds is determined from equation 12.5 and the egg
removal from equation 11.12.
If the effluent is to be used for unrestricted irrigation, then by the time the
E coli numbers are reduced to ≤1000 per 100 ml, the eggs will have been
reduced to <<1 (even <0.1) egg/l – but this should always be checked.
BOD removal in maturation ponds is much slower than in facultative ponds
(Table 12.1). Equation 5.7 for unfiltered BOD can be used with a value of k1
Maturation Ponds 149
of ~0.05 day–1 for temperatures of 15–25°C, as there is little variation of BOD
removal in maturation ponds with temperature. The non-algal (ie filtered)
BOD is then estimated from equation 11.8 with Fna = 0.1 (ie assuming that
90 per cent of the BOD is algal).
Total nitrogen
There is no total nitrogen removal in anaerobic ponds, only nitrogen
transformation with some of the organic nitrogen (principally urea and amino
acids) being converted to ammonia (a process called ‘ammonification’).
Nitrogen removal occurs in facultative and maturation ponds, mainly through
the incorporation of ammonia into algal cells. Reed (1985) gives the following
equation for total nitrogen removal in facultative and maturation ponds:
Ce = Ciexp{– [0.0064(1.039)T–20][θ + 60.6(pH – 6.6)]}
where Ce and Ci are the effluent and influent total nitrogen concentrations,
respectively, mg N/l. If the pond pH is unknown, it may be estimated from the
pH = 7.3exp(0.0005A)
where A is the alkalinity, mg CaCO3/l.
The range of temperature used in the derivation of equation 12.12 was
1–28°C, so it can be confidently used in developing countries. It has direct
application in the design of wastewater-fed fishponds (Chapter 22).
Ammonia is principally removed in facultative and maturation ponds by
incorporation into algal biomass, although at high pH it may also be lost by
volatilization to the atmosphere. Pano and Middlebrooks (1982) give two
equations for ammonia removal in facultative and maturation ponds:
for T ≤20°C:
Ce = Ci/{1 + [(A/Q)(0.0038 + 0.000134T)exp((1.041 + 0.044T)(pH – 6.6))]}
150 Domestic Wastewater Treatment in Developing Countries
for T > 20°C:
Ce = Ci/{1 + [5.035 x 10–3(A/Q)][exp(1.540 x (pH – 6.6))]}
where Ce and Ci are the effluent and influent ammonia concentrations,
respectively, mg N/l. The pH is estimated from equation 12.13.
Both these equations are in fact versions of equation 5.7. Since A/Q = θ/D,
k1 (here the first-order rate constant for ammonia-N removal) is given for
equation 12.14 by:
k1 = (1/D)(0.0038 + 0.000134T)exp[(1.041 + 0.044T)(pH – 6.6)]
Equation 12.14 or 12.15 is also used in the design of wastewater-fed fishponds
(Chapter 22).
Recent research on ammonia removal in WSP in northeast Brazil
Pano and Middlebrooks (1982) derived their equations for ammonia removal
in facultative ponds which were receiving BOD loads of only 10–40 kg/ha day,
much less than those used in warm climates (Chapter 11). Silva et al (1995*)
investigated nitrogen removal in WSP in northeast Brazil: they found that the
Pano and Middlebrooks model was satisfactory for ammonia removal in
facultative and primary maturation ponds, but not for removal in secondary
and tertiary maturation ponds. Results for all facultative and maturation
ponds showed that ammonia removal at 25–27°C was very well predicted by
the equation:
Cc = Ci/[1 + 8.65 x 10–3(A/Q)exp(1.727(pH – 6.6))]
where Cc is the in-pond column-sample ammonia concentration, mg N/l (see
Chapter 14 for details of in-pond column sampling). The range of BOD
loadings on the ponds used to derive equation 12.17 was 20–220 kg/ha day.
Ammonia removal in a five-pond series with an overall retention time of 19
days was 91 per cent. Silva et al (1995*) also give the following equation for
TKN removal in facultative and maturation ponds (TKN is total Kjeldahl
nitrogen, which is organic N + ammonia N):
Cc = [(0.19/λTKN
) – 0.063]–1
where Cc is the in-pond TKN concentration, mg N/l; and λTKN
is the TKN
loading, kg N/ha/day (range 20–170 kg N/ha day).
Maturation Ponds 151
Nitrification, the obligately aerobic autotrophic oxidation of ammonia to
nitrite and then nitrate (Chapter 3), does not appear to be common in WSP;
nitrifying bacterial populations are small (but they are present). However,
strong claims for nitrification occurring in the WSP at the Western Treatment
Plant in Melbourne, Australia (Chapter 9) have been made by Hurse and
Connor (1999*; see also Lai and Lam, 1997*), with up to 107 nitrifying
bacteria per millilitre in the water column in Ponds 6–10 of the 10-pond system
which had an average overall retention time of 80 days; the annual mean
temperature variation was 8–23°C. This may explain why WSP in warm
climates do not nitrify – overall retention times are commonly <<40 days.
Phosphorus removal in facultative and maturation ponds occurs mainly
through precipitation as calcium hydroxyapatite at pH >9. However, overall P
removal in a series of WSP is often only ~50 per cent, with effluent
concentrations usually >3 mg P/l. If lower concentrations are required by the
regulator, in-pond dosing with aluminium sulphate (‘alum’) or ferric chloride
can be effective in reducing P levels from up to 15 mg/l to ~1 mg/l, without
causing significant sludge accumulation (Surampalli et al, 1995*).
When WSP effluents are discharged to inland waters, it may be necessary to
remove the algal suspended solids so that the effluent can comply with the
required quality for suspended solids. In the European Union this is ≤150 mg
SS/l for WSP effluents (Chapter 4), and so algal removal would not be
necessary. However, if a requirement of <50 mg/l is set, then it may be
necessary to remove the algal solids (but see Table 12.1 which shows a final
suspended solids concentration of 45 mg/l). If algal solids removal is necessary,
it is best done in rock filters (Middlebrooks, 1988, 1995*; Environmental
Protection Agency, 2002*; Neder et al, 2002*), although a final deep
maturation pond may achieve the same effect as light penetration is lower, and
algal solids are correspondingly lower, at greater depths.
Rock filters consist of a bed of rock which is full of pond effluent and
within which the algae settle out as the effluent flows horizontally through the
filter. The algae decompose releasing nutrients which are utilized by bacteria
growing on the surface of the rocks. Rock filters can be in-pond filters within
the final maturation pond, but it is operationally better (for ease of
maintenance) to have them as separate units following the final maturation
pond. Rock size is ~50–100 mm (~100–200 mm is also used), with a bed depth
of 0.5–1 m. The rocks should extend at least 100 mm above the water level in
the filter so as to avoid mosquito breeding and odours due to cyanobacteria
(blue green algae) growing on wet rocks exposed to sunlight.
152 Domestic Wastewater Treatment in Developing Countries
Rock filter design
Rock filters are designed on the basis of hydraulic loading rate, expressed as
m3 of pond effluent per m3 gross rock filter volume (ie ignoring the space
occupied by the rocks) per day; its units are thus day–1. Early work in the US
produced the following equations for percentage BOD and SS removal
(Swanson and Williamson, 1980; see also Johnson and Mara, 2002):
RBOD = 72 – 109(HLR)
RSS = 97 – 137(HLR)
where HLR is the hydraulic loading rate, day–1.
The range of HLR investigated by Swanson and Williamson was 0.06–0.34
day–1, which is low for warm climates. In northeast Brazil, Mara et al (2001b*)
found that rock filters receiving primary maturation pond effluent at an HLR
of 1 day–1 achieved BOD and SS removals of 46 and 62 per cent at 25°C,
respectively, but at an HLR of 2 day–1 the removals were only 14 and 53 per
cent. In Amman, Jordan, Saidam et al (1995*) used rock filters to treat the
final pond effluent; they found that rock filters of 20–230 mm ‘wadi gravel’
achieved BOD and SS removals of 49 and 46 per cent at an HLR of 0.27 day–1
at 25°C, but at an HLR of 0.39 day–1 the removals of both were reduced to 41
per cent. The following equation was derived for SS removal:
SSe = 0.88SSi – 1.92θT
where SSe and SSi are the effluent and influent suspended solids respectively,
mg/l; θ is the rock filter retention time, days; and T is the temperature, °C. The
retention time is given by:
θ = εVrf /Q
where ε is the porosity of the rock medium (around 0.4), and Vrf is the gross
rock filter volume, m3.
Design a maturation pond system to treat the effluent from the secondary
facultative pond calculated in Chapter 11 to produce (a) ≤1 egg/l and ≤0.1
egg/l/, and (b) ≤1000 E coli/100 ml by the methods of Marais (1974) and von
Sperling (1999*, 2002*, 2003*).
Maturation Ponds 153
Helminth eggs
A single maturation pond with the minimum retention time of 3 days achieves
a 90 per cent egg reduction (equation 11.12). Therefore:
Ee = 9(1 – 0.90)
= 0.9 egg/l
This complies with the World Health Organization’s (1989*, 2004*) guideline
of ≤1 egg/l. For compliance with ≤0.1 egg/l when children under 15 years are
exposed, a second 3-day maturation pond would be required:
Ee = 0. 9(1 – 0.90)
= 0.09 egg/l
The surface BOD loading on the first 3-day maturation pond is given by
equation 11.1 as:
λs = 10LiQ/Am
= 10LiDm/θm
= 10 x 60 x 1/3
= 200 kg/ha day
This is less than 70 per cent of the loading of 350 kg/ha day on the secondary
facultative pond, and therefore satisfactory.
E coli removal – Marais’ method
The E coli count in the effluent of the secondary facultative pond is given by
equations 12.3 and 12.2:
kB(T) = 2.6(1.19)T–20 = 6.2 day–1 for T = 25°C
Ne = Ni/(1 + kB(T)θa)(1 + kB(T)θf)
= 5 x 107/[1 + (6.2 x 1)][1 + (6.2 x 4)]
= 2.7 x 105 per 100 ml
The E coli count in the effluent of the first 3-day maturation pond is given by
equation 12.1 as:
154 Domestic Wastewater Treatment in Developing Countries
Ne = 2.7 x 105/[1 + (6.2 x 3)]
= 1.4 x 104 per 100 ml
This is <105/100 ml, the guideline value for restricted irrigation (Chapter 21).
The second 3-day maturation pond reduces the E coli count to:
Ne = 1.4 x 104/[1 + (6.2 x 3)]
= 700/100 ml
This is <1000/100 ml, the guideline value for unrestricted irrigation.
Thus to produce an effluent for unrestricted irrigation requires – in this
case and with Marais’ model – a total retention time of only 11 days (1 day in
the anaerobic pond, 4 days in the secondary facultative pond and 3 days in
each of the two maturation ponds).
The area of the first maturation pond is:
Am1 = 2Qiθm/(2Dm + 0.001eθm)
where Qi is the effluent flow from the secondary facultative pond, = 9868
Am1 = 2 x 9868 x 3/[(2 x 1) + (0.001 x 5 x3)]
= 29,380 m2
Qe = 9868 – (0.001 x 5 x 29,380)
= 9721 m3/day
The area of the second maturation pond is:
Am2 = 2 x 9721 x 3/[(2 x 1) + (0.001 x 5 x 3)]
= 28,950 m2
Qe = 9721– (0.001 x 5 x 29 380)
= 9576 m3/day
Thus the total area required for treatment to ≤1000 E coli/100 ml is:
Anaerobic pond
Secondary facultative pond
First maturation pond
3340 m2
26, 500 m2
29, 380 m2
Maturation Ponds 155
Second maturation pond
Total pond area
28, 950 m2
88, 170 m2
For the population served of 100,000, this equals 0.88 m2/person, or 88
ha/million people. Allowing for embankments, etc, the overall area would be
~(1.25 x 88), = ~110 ha/million people.
The above design example is very straightforward. To illustrate the general
design procedure in greater detail, consider a design temperature of 15°C
(rather than 25°C).
Anaerobic pond: the retention time is now 1.5 days and the BOD removal
50 per cent.
Secondary facultative pond: the design surface BOD loading is 170 kg/ha
day; the retention time is now 13.5 days and the effluent BOD is 72 mg/l.
Primary maturation pond: the retention time in the first maturation pond
is given by equation 12.5 as:
Qm1 = 10Le(f)Dm/0.75λs(f)
= (10 x 72 x 1)/(0.75 x 170)
= 5.7 days
E coli removal: the value of kB at 15°C is given by equation 12.2 as 1.09
day–1; therefore the number of E coli/100 ml of the primary maturation
pond effluent, as given by equation 12.3 is:
Ne = Ni/(1 + kBθa)(1 + kBθf)(1 + kBθm1)
= 5 x 107/[1 + (1.09 x 1.5)][1 + (1.09 x 13.5)][1 + (1.09 x 5.7)]
= 1.7 x 105 per 100 ml
Subsequent maturation ponds: the following version of equation 12.7 is
θm = [(Ni/Ne)1/n – 1]/kB
This equation, with Ni = 1.7 x 105/100 ml and Ne = 1000/100 ml, is now solved
for n = 1, 2, etc until the calculated value of θm is <3 days:
for n = 1, θm = 155 days
for n = 2, θm = 11 days
for n = 3, θm = 4.2 days
for n = 4, θm = 2.4 days – thus ñ = 4
156 Domestic Wastewater Treatment in Developing Countries
Consider, therefore, the combinations of 3 ponds @ 4.2 days and 4 ponds @ 3
min). Choose the former, as it is only 0.6 days more but will minimize
days (θm
embankment construction. The areas of all four maturation ponds (ie including
the primary maturation pond) are calculated as shown above.
E coli removal – von Sperling’s method
• Anaerobic pond: the E coli count in the anaerobic pond effluent is given
by equations 12.1 and 12.8 as:
kB(25) = 2.0(1.07)5 = 2.8 day–1
Ne = 5 x 107/[1 +(2.8 x 1)] = 1.3 x 107/100 ml
Secondary facultative pond: assume a retention time of 4 days, a lengthto-breadth ratio of 3 to 1 and a depth of 1.5 m. Equations 5.8 and
5.13–5.15 give:
kB(20) = 0.92(1.5)–0.88(4)–0.33 = 0.41 day–1
kB(25) = 0.41(1.07)5 = 0.56 day–1
δ = 0.33
a = [1 + (4 x 0.56 x 4 x 0.33)]0.5 = 1.99
Ne = [1.3 x 107][(4 x 1.99)/(1 + 1.99)2]exp[(1 – 1.99)/(2 x 0.33)]
= 2.6 x 106 per 100 ml
Maturation ponds: take θm = θmmin = 3 days and a depth of 1 m; assume
that the ponds are baffled to give a length-to-breadth ratio of 10 to 1.
Equations 5.8 and 5.13–5.15 give for each pond:
kB(20) = 0.92(1)–0.88(3)–0.33 = 0.64 day–1
kB(25) = 0.64(1.07)5 = 0.90 day–1
δ = 0.1
a = [1 + (4 x 0.90 x 3 x 0.1)]0.5 = 1.44
Ne = Ni[(4 x 1.44)/(1 + 1.44)2]exp[(1 – 1.44)/(2 x 0.1)]
= Ni x 0.107
Maturation Ponds 157
For the effluent from the fourth maturation pond:
Ne = 2.6 x 106 x (0.107)4 = 360/100 ml
Thus, for unrestricted irrigation, the WSP series comprises a 1-day anaerobic
pond, a 4-day secondary facultative pond and four 3-day maturation ponds,
giving a total retention time of 17 days.
The von Sperling design has a total retention time of 17 days and the Marais
design 11 days. This discrepancy can be explained (at least partially) by the
fact that Marais’ equation for kB(T) was derived from a single WSP series close
to its optimal loading, whereas von Sperling’s equations were derived from
many WSP systems which were not all optimally loaded. In practice, of course,
it would be unusual for WSP systems to be at their optimal (ie correct design)
loading – for most of the time they are either underloaded or, more commonly,
overloaded. Von Sperling’s method may therefore reflect actual (as opposed to
optimal) faecal coliform removal in a series of WSP.
Physical design of WSP
The WSP process design prepared as described in Chapters 10–12 must be
‘translated’ into a physical design. Actual pond dimensions, consistent with
the available site, must be calculated; embankments and pond inlet and outlet
structures must be designed; and decisions taken regarding preliminary
treatment (Chapter 8), how many parallel pond series to have, and whether or
not to line the ponds. By-pass pipework, security fences and notices are
generally required, and operator facilities should be provided.
The physical design of WSP must be carefully done: it is at least as
important as process design and can significantly affect treatment efficiency
(Bernhard and Degoutte, 1990; Drakides and Trotouin, 1991). Advice on the
preparation of engineering drawings for WSP is given by the Agency for
International Development (1982*).
Ponds should be located at least 200 m (preferably 500 m) downwind from
the community they serve and away from any likely area of future expansion;
this is to discourage people from visiting the ponds. Odour release, even from
anaerobic ponds, is most unlikely to be a problem in a well-designed and
properly maintained system, but the public may need assurance about this at
the planning stage, and a minimum distance of 200–500 m normally allays
any fears (in some parts of the developing world people live immediately
adjacent to WSP, and they generally like the location of their house – it has a
nice view!).
There should be vehicular access to and around the ponds and, in order to
minimize earthworks, the site should be flat or gently sloping. The soil must
also be suitable. Ponds should not be located within 2 km of airports, as birds
attracted to the ponds constitute a risk to air navigation.
Geotechnical aspects of WSP design are very important. In France, for
example, a third of the WSP systems that malfunction do so because of
Physical design of WSP 159
geotechnical problems which could have been avoided at the design stage
(Bernhard and Kirchgessner, 1987). Poor geotechnical design is also common
in Mexico (Mantilla et al, 2002) and presumably elsewhere.
The principal objectives of a geotechnical investigation are to ensure
correct embankment design and to determine whether the soil is insufficiently
impermeable to require the pond to be lined. The maximum height of the
groundwater table should be determined, and the following properties of the
soil at the proposed pond location must be measured:
particle size distribution,
maximum dry density and optimum moisture content (modified Proctor
Atterberg limits,
organic content, and
coefficient of permeability.
At least one soil sample should be taken per hectare and the samples should be
as undisturbed as possible. They should be representative of the soil profile to
a depth 1 m greater than the envisaged pond depth.
Organic and plastic soils and medium-to-coarse sands are not suitable for
embankment construction. If there is no suitable local soil with which at least
a stable and impermeable embankment core can be formed, it must be brought
to the site at extra cost and the local soil, if suitable, used for the embankment
Ideally, of course, embankments should be constructed from the soil
excavated from the site, and there should be a balance between cut and fill,
although it is worth noting that ponds constructed completely in cut may be a
cheaper alternative, especially if embankment construction costs are high. The
soil used for embankment construction should be compacted in 150–250 mm
layers to 90 per cent of the maximum dry density as determined by the
modified Proctor test. Shrinkage of the soil occurs during compaction (by
10–30 per cent) and excavation estimates must take this into account. After
compaction, the soil should have an in situ coefficient of permeability of <10–7
m/s. Wherever possible and particularly at large pond installations,
embankment design should allow for vehicular access to facilitate
Embankment slopes are commonly 1 to 3 internally and 1 to 1.5–2
externally. Steeper slopes may be used if the soil is suitable; slope stability
should be ascertained according to standard soil mechanics procedures for
small earth dams. Embankments should be planted with grass to increase
stability; a slow-growing rhizomatous species (eg Bermuda grass) should be
used to minimize maintenance.
External embankments should be protected from stormwater erosion by
providing adequate drainage. Internal embankments require protection against
erosion by wave action, and this is best achieved by lean concrete cast in situ,
precast concrete slabs or stone rip-rap at top water level (see Figures
160 Domestic Wastewater Treatment in Developing Countries
Figure 13.1 Embankment Protection by Concrete Cast in situ
13.1–13.3). Such protection also prevents vegetation from growing down the
embankment into the pond, so preventing the development of a shaded habitat
suitable for mosquito or snail breeding.
Physical design of WSP 161
Figure 13.2 Embankment Protection by Precast Concrete Slabs
162 Domestic Wastewater Treatment in Developing Countries
Figure 13.3 Embankment Protection by Stone Rip-rap
Ponds should be lined if the soil is too permeable. As a general guide the
following interpretations can be given to the values obtained for the in situ
coefficient of permeability, k:
k >10–6 m/s: the soil is too permeable and the ponds must be lined,
k <10–7 m/s: some seepage may occur but not sufficiently to prevent the
ponds from filling,
k <10–8 m/s: the ponds will seal naturally, and
k <10–9 m/s: there is no risk of groundwater contamination (if k >10–9 m/s
and the groundwater is used for potable supplies, a detailed
hydrogeological investigation will be required).
A variety of lining materials can be used when k is >10–6 m/s; local availability
and costs will determine which should be used. Plastic liners are commonly
used (Figures 13.4 and 13.5); alternatively a 300-mm thick clay liner can be
used. Advice on pond lining is given by Environment Protection Agency
Physical design of WSP 163
Figure 13.4 Anaerobic Pond Lined with an Impermeable Plastic Membrane
There has been little rigorous work done on determining optimal pond shapes.
The most common shape is rectangular, although there is much variation in
the length-to-breadth ratio. Clearly, the optimal pond geometry, which
includes not only the shape of the pond but also the relative positions of its
inlet and outlet, is that which minimizes hydraulic short-circuiting (Persson,
In general, anaerobic and primary facultative ponds should be rectangular,
with length-to-breadth ratios of 2–3 to 1 so as to avoid sludge banks forming
near the inlet. However, the geometry of secondary facultative and maturation
ponds is less important than previously thought (Pearson et al, 1995*); they
can have higher length-to-breadth ratios (up to 10 to 1) so that they better
approximate plug flow conditions. Ponds do not need to be strictly
rectangular; they can be gently curved if necessary or if desired for aesthetic
reasons. A single inlet and outlet are usually sufficient, and these should be
located in diagonally opposite corners of the pond (the inlet should not
discharge centrally in the pond as this maximizes hydraulic short-circuiting).
To minimize hydraulic short-circuiting, the inlet should be located such that
the wastewater flows into the pond against the prevailing wind.
Baffles should only be used with caution. In facultative ponds, when baffles
are needed because the site geometry is such that it is not possible to locate the
inlet and outlet in diagonally opposite corners, care must be taken in locating
164 Domestic Wastewater Treatment in Developing Countries
Trench cut by trenching
machine – insert lining
backfill and compact
12” to 16”
1% slope
Top of slope
6” min
Stable, compact soil
or existing concrete,
dunite or asphalt concrete
Source: Environmental Protection Agency (1983)
Figure 13.5 Anchoring the Pond Liner at the Top of the Embankment
the baffle(s) in order to avoid too high a BOD loading in the inlet zone (and
the consequent possible risk of odour release). In maturation ponds baffling is
advantageous as it helps to maintain the surface zone of high pH, which
facilitates the removal of faecal bacteria (Pearson et al, 1995*; Lloyd et al,
2003*). Recent research into pond hydraulics by Shilton (2001*) and Shilton
and Harrison (2003a*), which involved computer modelling, laboratory
investigations and testing on full-scale ponds, has demonstrated that the energy
(ie momentum) of the pond influent flow is more important than wind effects
in determining flow patterns within the pond and hence the degree of hydraulic
short-circuiting induced. It was found that short ‘stub’ baffles near the inlet
and outlet were as effective as two long baffles, each extending 70 per cent of
the pond width, in reducing E coli levels. Persson (2000*) found that a small
island (with 2 per cent of the pond area) located near the inlet achieved very
little short-circuiting, with the effective pond volume being 96 per cent of the
actual pond volume. Persson’s inlet island and Shilton and Harrison’s inlet
stub baffle are essentially the same hydraulic device (as is the inlet scum box
shown in Figure 13.8). Pond hydraulics are discussed in detail by Shilton and
Harrison (2003b*), which should be consulted for further details.
The areas calculated by the process design procedures described in
Chapters 10–12 are mid-depth areas, and the dimensions calculated from them
are thus mid-depth dimensions. These need to be corrected for the slope of the
embankment, as shown in Figure 13.6. A more precise method is advisable for
anaerobic ponds, as these are relatively small; the following formula is used
(Environmental Protection Agency, 1983):
Physical design of WSP 165
L – nD
Figure 13.6 Calculation of Top and Bottom Pond Dimensions from
those Based on Mid-depth
Va = [(LW) + (L – 2sD) (W – 2sD) + 4(L – sD)] [D/6]
where Va is the anaerobic pond volume, m3; L is the pond length at top water
level (TWL), m; W is the pond width at TWL, m; s is the horizontal slope
factor (ie a slope of 1 in s); and D is the pond liquid depth, m. With the
substitution of L as nW, based on a length-to-breadth ratio of n to 1, equation
13.1 becomes a simple quadratic in W.
The dimensions and levels that the contractor needs to know are those of
the base and the top of the embankment; the latter includes the effect of the
freeboard. The minimum freeboard that should be provided is decided on the
basis of preventing wind-induced waves from overtopping the embankment.
For small ponds (<1 ha in area) 0.5 m freeboard should be provided; for ponds
between 1 ha and 3 ha, the freeboard should be 0.5–1 m, depending on site
considerations. For larger ponds the freeboard may be calculated from the
equation (Oswald, 1975):
F = (log10A)1/2 – 1
where F is the freeboard, m; and A is the pond area at TWL, m2.
Pond liquid depths are commonly in the following ranges:
anaerobic ponds:
facultative ponds:
maturation ponds:
2–5 m
1–2 m
1–1.5 m
The depth chosen for any particular pond depends on site considerations (eg
the presence of shallow rock, minimization of earthworks). The depth of
facultative and maturation ponds should be ≥1 m so as to avoid vegetation
growing up from the pond base, with the consequent hazard of mosquito and
snail breeding.
166 Domestic Wastewater Treatment in Developing Countries
At WSP systems serving >10,000 people, it is often sensible (so as to
increase operational flexibility) to have two or more series of ponds in parallel.
The available site topography may in any case necessitate such a subdivision,
even for smaller systems. Usually the series are equal, that is to say they receive
the same flow, and arrangements for splitting the raw wastewater flow into
equal parts after preliminary treatment must be made (see Stalzer and von der
Emde, 1972). This is best done by providing weir penstocks ahead of each
It is very important to divide the total area for each type of pond, as
calculated from the process design equations in Chapters 10–12, in parallel
and not in series. Thus, if it is decided to have n series of ponds at a particular
site, each anaerobic pond has an area of Aa/n, each facultative pond one of
Af/n, and so on. (If they are divided in series, then the loading on the first pond
is n times higher than it should be and pond failure is guaranteed.)
There is a wide variety of designs for the inlet and outlet structures, and
provided they follow certain basic concepts, their precise design is relatively
unimportant. First, they should be simple and inexpensive. While this should
be self-evident, it is all too common to see unnecessarily complex and
expensive structures. Second, they should permit samples of the pond effluent
to be taken with ease. The inlet to anaerobic and primary facultative ponds
Scum box
Note: The scum box retains most of the floating solids, so easing pond maintenance; it also
improves pond hydraulics.
Source: ABLB and CTGREF (1979)
Figure 13.7 Inlet Structure for Anaerobic and Primary Facultative Ponds
Physical design of WSP 167
Figure 13.8 Inlet Structure on a Facultative Pond with Integral Scum Box
should discharge well below the liquid level so as to minimize short-circuiting
(especially in deep anaerobic ponds) and reduce the quantity of scum (which is
important in facultative ponds). Inlets to secondary facultative and maturation
ponds should also discharge below the liquid level, preferably at mid-depth in
order to reduce the possibility of short-circuiting. Some simple inlet designs
are shown in Figures 13.7–13.9.
Single inlets and outlets in diagonally opposite corners are best.
Occasionally multi-inlets and multi-outlets are used in the mistaken belief that
they improve pond hydraulics. However, what often happens is that, due to
poor construction, one of the outlets settles and its discharge level is then lower
than those of the others, with the result that all the pond effluent discharges
through this outlet and the others are left, literally, high and dry.
All outlets should be protected against the discharge of scum by the
provision of a scum guard. The take-off level for the effluent, which is
controlled by the scum guard depth, is important as it has a significant
influence on effluent quality. In facultative ponds, the scum guard should
extend just below the maximum depth of the algal band when the pond is
stratified so as to minimize the daily quantity of algae, and hence BOD, leaving
the pond. In anaerobic and maturation ponds, where algal banding is
irrelevant, the take-off should be nearer the surface: in anaerobic ponds it
should be well above the maximum depth of sludge but below any surface
crust, and in maturation ponds it should be close to the surface to give the best
possible microbiological quality. The following effluent take-off levels are
168 Domestic Wastewater Treatment in Developing Countries
150 mm concrete surround
where cover less than 600 mm
1 m wide concrete slab
for erosion protection
Base of pond
Note: This would receive the discharge from the outlet structure shown in Figure 13.10.
Figure 13.9 Inlet Structure for Secondary Facultative and Maturation Ponds
Anaerobic ponds:
Facultative ponds:
Maturation ponds:
300 mm
600 mm
50 mm
The installation of a variable height scum guard is recommended, since it
permits the optimal take-off level to be set once the pond is operating.
A simple outlet weir structure is shown in Figure 13.10. The following
formula should be used to determine the head over the weir and so, knowing
the pond depth, the required height of the weir above the pond base can be
q = 0.0567h3/2
where q is the flow per metre length of weir, l/s; and h is the head of water
above weir, mm.
The outlet from the final pond in a series should discharge into a simple
flow-measuring device such as a triangular or rectangular notch. Since the flow
into the first pond is also measured, this permits the rate of evaporation and
seepage to be calculated or, if evaporation is measured separately, the rate of
It is necessary to by-pass anaerobic ponds so that facultative ponds may be
commissioned first and also during desludging operations. Figure 13.11 shows
a by-pass arrangement for two series of WSP in parallel.
Physical design of WSP 169
Concrete weir
scum guard
Base of
Note: The weir length is calculated from equation 13.3. The discharge pipe would connect with
the inlet structure shown in Figure 13.9. The concrete scum guard depth should be appropriate
for the type of pond it is in (here, it is 600 mm, suitable for facultative ponds); as an alternative a
variable-depth wooden scum guard may be used.
Figure 13.10 Outlet Weir Structure
Covering an anaerobic pond allows the biogas to be collected and used
(commonly for electricity generation), and it also minimizes any potential
odour problem (Figures 13.12 and 16.3). DeGarie et al (2000*) describe a
floating composite cover for the anaerobic section of the first pond in the large
WSP systems at the Western Treatment Plant in Melbourne, Australia, each of
which receives a wastewater flow of 120,000 m3/day (Chapter 9). The cover is
composed of three layers: a high-tensile-strength UV-resistant geomembrane
for biogas recovery at the top, a 12.5-mm polyfoam insulation and flotation
layer in the middle which is welded to a base layer of high-density
polyethylene. The cover measures 171 x 200 m (ie an area of 3.4 ha). The
biogas collected from the ponds is used to generate 6000 kW of electricity
8–16 hours per day, 365 days per year, which is worth around Aus$1.8 million
(US$1.1 million) per year.
Only anaerobic ponds at large WSP systems are suitable for biogas
collection and energy generation. Methane, which comprises around 70 per
cent of the biogas, is a powerful greenhouse gas, so its utilization for energy
generation at large WSP systems is, environmentally, a very sensible option.
In desert areas a treebelt should be provided to prevent wind-blown sand from
being deposited in the ponds. Treebelts may also be desired for aesthetic
170 Domestic Wastewater Treatment in Developing Countries
to M1
to M2
Note: During normal operations, gate G3 is closed and the others open; to by-pass the anaerobic
ponds gate G3 is opened and the others closed
Figure 13.11 By-pass Pipework for Anaerobic Ponds
reasons if the WSP site is close to human habitation. They should be planted
upwind of the WSP and comprise up to five rows, as follows (from the upwind
1–2 rows of mixed shrubs such as Latana, Hibiscus and Nerium oleander
(none of which is eaten by goats);
1–2 rows of Casuarina trees; and
1 row of a mixture of taller trees such as Poinciana regia (flame trees),
Tipuana tipu, Khaya senegalensis and Albizia lebbech.
Figure 13.12 Covered Anaerobic Pond at the Western Treatment Plant,
Melbourne, Australia
Physical design of WSP 171
Local botanists will be able to advise on which species are most appropriate
(see also Wickers et al, 1985); those given above are suitable for use in North
Africa. Such a treebelt is around 40–60 m wide. It should be irrigated with
final effluent.
If food trees (for example, olive trees) are also grown, then sale of the
produce (either directly or by concession) can contribute significantly to
operation and maintenance (O&M) costs. For example, at the large WSP
system at Al Samra, serving the cities of Amman and Zarqa (combined
population: 2.6 million) in Jordan, over 1.5 million trees were planted around
the site (which is a desert), and they are irrigated with 2–3 per cent of the final
effluent. Some 60,000 mature olive trees are leased to a local farmer who pays
US$12,000 per year for this, which contributes significantly to the O&M costs
(Figure 13.13).
Ponds (other than at very remote locations) should be surrounded by a chainlink fence with gates which should be kept padlocked. Warning notices in the
appropriate local language(s) (Figure 13.14), advising that the ponds are a
wastewater treatment facility and therefore potentially hazardous to health,
are essential to discourage people from visiting the ponds, which if properly
maintained can appear as pleasant, inviting bodies of water. Children are
especially at risk, as they may be tempted to swim or play in the ponds.
In many parts of the developing world particular attention must be paid to
keeping wild animals away from the ponds, especially hippopotamuses and
crocodiles. Low-voltage electric fences (12 V DC) are effective in keeping hippos
out (and this is important: an operator at the Dandora WSP in Nairobi was
killed by a hippo in 1989). To keep out crocodiles, the chain link fence should
extend 50 cm below ground level; this section and the first 50 cm above ground
level should be in small-aperture chain-link fencing to keep out baby crocodiles.
Figure 13.13 Partial View of the Al Samra WSP Showing some of the Olive
Trees, the Produce from which helps Pay for the System’s O&M
172 Domestic Wastewater Treatment in Developing Countries
Figure 13.14 Fence and Warning Notice in English and Kiswahili at a Pond
Site in Nairobi, Kenya
The facilities to be provided for the team of pond operators depend partly on
their number (Chapter 14), but would normally include the following:
a first-aid kit (which should include a snake-bite kit);
strategically placed lifebuoys;
a wash-hand basin and toilet; and
storage space for protective clothing, grass-cutting and scum-removal
equipment, screen rakes and other tools, a sampling boat (if provided) and
Physical design of WSP 173
With the exception of the lifebuoys, these can be accommodated in a simple
building. This can also house, if required, sample bottles and, if electricity is
available, a refrigerator for sample storage. Laboratory facilities, offices and a
telephone may also be provided at large installations. There should be
vehicular access and space for parking.
Prior to upgrading or extending a WSP system, its performance should be
evaluated as described in Chapter 15, as this will generally permit the correct
decision about how to upgrade and/or extend the system to be made.
A number of strategies can be used to upgrade and extend WSP systems
(Environmental Protection Agency, 1977). In addition to any rehabilitation
measures needed (Chapter 14), these include:
the provision of anaerobic ponds;
the provision of additional maturation ponds;
the provision of one or more additional series of ponds; and/or
the alteration of pond sizes and configuration – for example, the removal
of an embankment between two ponds to create a larger one.
Original design:
10 days
3 days
3 days
Upgraded system:
1 day
5 days
3 days
3.7 days
3.7 days
Note: The embankment between the original maturation ponds becomes a baffle in the upgraded
first maturation pond. The total retention time is increased only from 16 to 16.4 days. The
improvement in microbiological quality can be illustrated as follows, by using equation 12.3 with
Ni = 5 x 107 per 100 ml and kT = 2.6 day–1 (ie for 20°C):
Original design: Ne = 5 x 107/[(1 + (2.6 x 10)) (1 + (2.6 x 3))2]
= 24,000 per 100 ml
Upgraded system: Ne = 5 x 107/[(1 + (2.6 x 1)) (1 + (2.6 x 5)) (1 + (2.6 x 3)) (1 + (2.6 x 3.7))2]
= 11,000 per 1000 ml
Figure 13.15 Upgrading a WSP Series to Treat Twice the Original Flow
174 Domestic Wastewater Treatment in Developing Countries
Figure 13.15 shows how (1), (2) and (4) above can be combined to upgrade a
single series of WSP to receive twice its original design flow, with the
production of a higher quality effluent which meets the World Health
Organization’s (1989*, 2004*) guideline value for unrestricted irrigation.
Operation and Maintenance of WSP
WSP systems should preferably be commissioned at the beginning of the hot
season in order to establish as quickly as possible the necessary microbial
populations to effect waste stabilization. Prior to commissioning, all ponds
should be free from vegetation. Facultative ponds should be commissioned
before anaerobic ponds in order to avoid odour release when an anaerobic
pond effluent is discharged into an empty facultative pond. It is best to fill
facultative and maturation ponds first with river or lake water so as to permit
the gradual development of the algal and heterotrophic bacterial populations.
If such water is unavailable, facultative ponds should be filled with raw
wastewater and left for 3–4 weeks to allow the microbial populations to
develop; a little odour release may occur during this period.
Anaerobic ponds should be filled with raw wastewater and seeded, where
possible, with digesting sludge from, for example, local septic tanks. The ponds
should then be gradually loaded up to the design loading rate over the
following week (or month if the ponds are not seeded). Care should be taken
to maintain the pond pH above 7 to permit the development of methanogenic
bacteria, and it may be necessary during the first month or so to dose the pond
with lime or soda ash. If, due to an initially low rate of sewer connections in
newly sewered towns, the wastewater is weak or its flow low, it is best to bypass the anaerobic ponds (Figure 13.11) until the wastewater strength and
flow is such that a BOD loading of at least 50 g/m3 day can be applied to them.
(If it is planned to by-pass the anaerobic ponds during desludging (see below),
then the by-pass should be a permanent facility.)
The maintenance requirements of ponds are very simple, but they must be
carried out regularly. Otherwise, there may be serious odour, fly and mosquito
nuisance. Maintenance requirements and responsibilities must therefore be
clearly defined at the design stage so as to avoid problems later. Routine
maintenance tasks are as follows:
176 Domestic Wastewater Treatment in Developing Countries
the removal of screenings and grit from the preliminary treatment
processes (Chapter 8);
cutting the grass on the embankments and removing it so that it does not
fall into the pond (this is necessary to prevent the formation of mosquitobreeding habitats);
the removal of floating scum and floating macrophytes, such as Lemna
(duckweed), from the surface of facultative and maturation ponds (this is
required to maximize photosynthesis and surface reaeration and to prevent
fly and mosquito breeding);
spraying the scum on anaerobic ponds (which should not be removed as it
aids the treatment process), as necessary, with clean water or pond effluent,
or a suitable biodegradable larvicide, to prevent fly breeding;
the removal of any accumulated solids in the inlets and outlets;
repairing any damage to the embankments caused by rodents, rabbits or
other animals; and
repairing any damage to the external fences and gates.
The operators must be given precise instructions on the frequency with which
these tasks should be done, and their work should be constantly supervised.
The operators, after suitable training, may also be required to take samples for
subsequent laboratory analyses (Chapter 15).
Anaerobic ponds require desludging when they are around one-third full of
sludge (by volume). This occurs every n years where n is given by:
n = Va/3Ps
where Va is the volume of anaerobic pond, m3; P is the population served; and
s is the sludge accumulation rate, m3/person year.
A good design value for s in warm climates is 0.04 m3/person year. The
precise requirement for desludging can be determined by the ‘white towel’ test
of Malan (1964). White towelling material is wrapped along one-third of a
sufficiently long pole, which is then lowered vertically into the pond until it
reaches the pond bottom; it is then slowly withdrawn. The depth of the sludge
layer is clearly visible since some sludge particles will have been entrapped in
the towelling material (Figure 14.1). The sludge depth should be measured at
various points throughout the pond, away from the embankments, and its
mean depth calculated.
While an anaerobic pond must be desludged when it is one-third full of
sludge, which occurs every n years as given by equation 14.1, it may be
operationally easier to desludge it partially in the same month of every year. A
task that has to be done every February, for example, has a better chance of
Operation and Maintenance of WSP 177
Figure 14.1 Sludge Depth Measurement by the ‘White Towel’ Test
178 Domestic Wastewater Treatment in Developing Countries
Figure 14.2 Pond Desludging in Northern France using a Raft-mounted
Sludge Pump – detail: sludge suction head
Operation and Maintenance of WSP 179
being carried out than one to be done every n years (as it is often forgotten
exactly when the task was last undertaken).
Sludge can be readily removed by using a raft-mounted sludge pump.
These are commercially available (eg Brain Associates Ltd, Narberth,
Pembrokeshire SA67 7ES, UK), or they can be assembled locally: Figure 14.2
shows one such unit being used on a primary facultative pond in France. The
sludge is discharged into either adjacent sludge drying beds or tankers to
transport it to a landfill site, agricultural land or other suitable disposal
location. Although pond sludge has a better microbiological quality than that
from conventional treatment works, its disposal must be carried out in
accordance with any local regulations governing sludge disposal.
In order that the routine operation and maintenance (O&M) tasks can be
properly done, WSP installations must be adequately staffed. The level of
staffing depends on the type of inlet works (for example, mechanically raked
screens and proprietary grit removal units require an elctromechanical
technician, but manually raked screens and manually cleaned grit channels do
not), whether there are on-site laboratory facilities, and how the grass is cut
(manually or by mechanical mowers). Recommended staffing levels are given
in Table 14.1 for WSP systems serving populations up to 250,000; for larger
systems the number of staff should be increased pro rata.
Table 14.1 Recommended Staffing Levels for WSP Systems
Population served
Mechanical engineera
Laboratory technicianb
Assistant foreman
a Depends on amount of mechanical equipment used
b Depends on existence of laboratory facilities
c Depends on use of vehicle-towed lawn mowers, etc
d Depends on location and value of equipment used
Source: Arthur (1983*)
180 Domestic Wastewater Treatment in Developing Countries
If good pond O&M is not routinely carried out, the effects can become very
serious: odour release from both anaerobic and facultative ponds; fly breeding
in anaerobic ponds; floating macrophytes and/or emergent vegetation in
facultative and maturation ponds leading to mosquito breeding (Figure 14.3);
and in extreme cases the ponds can silt up and completely ‘disappear’.
Agunwamba (2001*) describes a very poorly maintained WSP system at a
university campus in West Africa: the wastewater flow and BOD had increased
over a 20-year period by 62 and 426 per cent, respectively, with the result that
odour release and mosquito breeding were occurring, and also flooding; some
livestock had drowned in the ponds, which were ~80 per cent full of sludge.
Examples, like this, of grossly poor pond maintenance are not, unfortunately,
uncommon (and they are often used to advocate that other, usually more
‘sophisticated’, wastewater treatment systems be used instead of WSP – but if
WSP cannot be maintained properly, what chance do other systems have?).
Pond rehabilitation is then necessary. This is achieved by a combination of
the following:
a complete overhaul (or redesign) of the inlet works, replacing any units
that cannot be satisfactorily repaired;
repairing or replacing any flow-measuring devices;
ensuring that any flow-splitting devices actually split the flow into the
required proportions;
desludging the anaerobic or primary facultative ponds, and any subsequent
ponds if necessary;
unblocking, repairing or replacing pond inlets and/or outlets;
relocating any improperly located inlets and/or outlets, so that they are in
diagonally opposite corners of each pond (or providing a baffle if
relocation is not feasible);
repairing, replacing or providing effluent scum guards;
preventing ‘surface streaming’ of the flow by discharging the influent at
removing scum and floating or emergent vegetation from the facultative
and maturation ponds;
checking embankment stability, and repairing, replacing or installing
embankment protection;
checking for excessive seepage (>10 per cent of inflow) and lining (or
relining) the ponds if necessary;
cutting the embankment grass; and
repairing or replacing any external fences and gates; fences may need to be
electrified to keep out wild and domestic animals.
As rehabilitation is commonly very expensive, good routine O&M is very
much more cost-effective.
Operation and Maintenance of WSP 181
Figure 14.3 A very Badly Neglected Facultative Pond in Eastern Africa
Monitoring and Evaluation of WSP
Once a WSP system has been commissioned, a routine monitoring programme
should be established so that the actual quality of its effluent can be
determined. This permits a regular assessment to be made of whether the
effluent is complying with local discharge or re-use standards. Moreover,
should a pond system suddenly fail or its effluent start to deteriorate, the
results of such a monitoring programme often give some insight into the cause
of the problem and so indicate what remedial action is required.
The evaluation of pond performance and behaviour, although a much
more complex procedure than the routine monitoring of effluent quality, is
nonetheless extremely useful as it provides information on how underloaded
or overloaded the system is, and thus by how much, if any, the loading on the
system can be safely increased as the community it serves expands, or whether
further ponds in parallel and/or in series are required (Chapter 13). It also
indicates how the design of future pond installations in the region can be
improved to take account of local conditions.
Effluent quality monitoring programmes should be simple and the minimum
required to provide reliable data. Two levels of effluent monitoring are
Level 1: representative samples of the final effluent should be taken
regularly (at least monthly) and analysed for those parameters for which
effluent discharge or re-use requirements exist.
Level 2: when Level 1 monitoring shows that a pond effluent is failing to
meet its discharge or re-use quality, a more detailed study is necessary.
Table 15.1 gives a list of the parameters whose values are required,
together with recommendations for the types of samples that should be
Since pond effluent quality shows a significant diurnal variation (although this
is less pronounced in anaerobic and maturation ponds than in facultative
Monitoring and Evaluation of WSP 183
ponds), 24-hour flow-weighted composite samples are preferable for most
parameters, although grab samples are necessary for some (pH, temperature
and E coli). Composite samples should be collected in one of the following
in an automatic sampler which takes grab samples every 1–2 hours, with
subsequent manual flow-weighting if this is not done automatically by the
by taking grab samples every 1–3 hours with subsequent manual flowweighting; or
by taking a column sample near the outlet of the final pond; this can be
done at any time of day and gives a good approximation (± ~20 per cent)
to the mean daily effluent quality (Pearson et al, 1987d).
Flow-weighting is used in order to determine more accurate estimates of mean
daily parameter values such as BOD and suspended solids. Grab samples are
taken every 1–3 hours for 24 hours, and the volume of each grab sample used
to make the 24-hour composite sample depends on the wastewater flow at the
time it was taken, for example, if at any time the flow were 10,000 m3/day,
then 100 ml of the grab sample taken at that time would be used to make the
24-hour composite; 150 ml would be used for a flow of 15,000 m3/day, and
230 ml for a flow of 23,000 m3/day, and so on. Thus the greater the flow, the
more ‘weight’ is given to the sample – hence the term ‘flow-weighting’.
A full evaluation of the performance of a WSP system is a time-consuming and
expensive process, and it requires experienced personnel to obtain and
interpret the data. However, it is the only means by which pond designs can be
optimized for local conditions. It is often, therefore, a highly cost-effective
exercise. The recommendations given below constitute a Level 3 monitoring
programme, and they are based on the guidelines for the minimum evaluation
of pond performance given by Pearson et al (1987e).
It is not intended that all pond installations be studied in this way, but only
one or two representative systems in each major climatic region. This level of
investigation is most likely to be beyond the capabilities of local organizations,
and it would need to be carried out by a state or national body, or by a
university under contract to such a body. This type of study is also necessary
when it is required to know how much additional loading a particular system
can receive before it is necessary to extend it (Chapter 13).
Samples should be taken and analysed on seven days over a seven-week
period at both the hottest and coldest times of the year. Samples are required of
the raw wastewater and of the effluent of each pond in the series and, so as to
take into account the weekly variation in influent and effluent quality, samples
should be collected on Monday in the first week, Tuesday in the second week
184 Domestic Wastewater Treatment in Developing Countries
Table 15.1 Parameters to be Determined for Level 2 Pond Effluent
Quality Monitoring
Sample typea
Suspended solids
E. coli
Total nitrogen
Total phosphorus
Electrical conductivity
Ca, Mg, Na
Helminth eggs
Measure both raw wastewater and final
effluent flows
Unfiltered samplesb
Unfiltered samplesb
Take two samples, one at 08.00 – 10.00 h
and the other at 14.00 – 16.00 h
Take sample between 08.00 and 10.00 h
Only when effluent being used (or being
assessed for use) for crop irrigation. Ca,
Mg and Na are required to calculate the
sodium absorption ratio (Chapter 22)
a C = 24-hour flow-weighted composite sample; G = grab sample
b Also on filtered samples if the discharge requirements are so expressed
and so on. Table 15.2 lists the parameters whose values are required. Generally
the analytical techniques described in the latest edition of Standard Methods
(American Public Health Association, currently 1998) are recommended,
although the modified Bailenger technique should be used for counting the
number of nematode eggs (Ayres and Mara, 1996*) and E coli is best counted
using modern selective media (such as chromogenic media, Chromagar, 2002*;
see also Environment Agency, 2002*).
Composite samples are necessary for most parameters, but grab samples
are required for temperature, pH and E coli, and samples of the entire pond
water column should be taken for algological analyses (chlorophyll a and algal
genera determination), using the pond column sampler shown in Figure 15.1.
Pond column samples should be taken from a boat or from a simple sampling
platform that extends beyond the embankment base (or from the outlet
structure if this extends sufficiently far into the pond). Data on at least daily
maximum and minimum air temperatures, rainfall and evaporation should be
obtained from the nearest meteorological station.
On each day that samples are taken, the mean mid-depth temperature of
each pond, which closely approximates the mean daily pond temperature,
should be determined by suspending a maximum-and-minimum thermometer
at the mid-depth of the pond at 8–9 am and reading it 24 hours later.
On one day during each sampling period, the depth of sludge in the
anaerobic and facultative ponds should be determined by the ‘white towel’ test
Monitoring and Evaluation of WSP 185
Table 15.2 Parameters to be Determined for the Minimum Evaluation of
WSP Performance
To be
determined fora
RW, all pond effluents
RW, all pond effluents
Suspended solids
RW, all pond effluents C
RW, all pond effluents G
All F and M pond contents
All F and M pond contents
RW, all pond effluents C
RW, A pond effluent, G, P
F pond contents or
depth profile
E. coli
Chlorophyll a
Algal genera
Total phosphorus
(mean daily)
RW, all pond effluents
RW, all ponds
Dissolved oxygenc
Depth profile in all F
and M ponds
Sludge depth
Ca, Mg and Na
Helminth eggs
A and F ponds
RW, all pond
Unfiltered and filtered
Unfiltered and filtered
Only if odour nuisance
present or facultative pond
effluent quality poor.
A depth profile is preferable
Use maximum–minimum
thermometers suspended in
RW flow and at mid-depth in
Measure at 08.00, 12.00
and 16.00 h on at least
three occasions
Use ‘white towel’ test
Only if effluent used or to be
used for crop irrigation. Ca,
Mg and Na required to
calculate the sodium
absorption ratio
(Chapter 22)
a RW, raw wastewater; FE, final effluent of pond series; A, anaerobic; F, facultative; M, maturation.
b C, 24 hour flow-weighted composite sample; G, grab sample taken when pond contents most
homogeneous; P, pond column sample.
c Measure depth profiles of pH and temperature at same times, if possible.
186 Domestic Wastewater Treatment in Developing Countries
Note: The overall length (here 1.7 m) may be altered as required. The design shown is a threepiece unit for ease of transportation, but this feature may be omitted. Alternative materials may be
used (eg PVC drainage pipe).
Figure 15.1 Details of Pond Column Sampler
(Figure 14.1). The sludge depth should be measured at various points
throughout the pond, away from the embankment base, and the mean depth
It is also useful to measure on at least one occasion during each sampling
season the diurnal variation in the vertical distribution of pH, dissolved oxygen
and temperature. Profiles should be obtained at 08.00, 12.00 and 16.00 h. If
submersible electrodes are not available, samples should be taken manually
every 15–20 cm.
It is advisable to store all data in a PC using a spreadsheet such as Excel, so
that simple data manipulations can be performed. From the data collected in
each sampling season (or month if sampling is done throughout the year),
mean values should be calculated for each parameter. Values, based on these
means, can then be calculated for:
the mean hydraulic retention time (= volume/flow) in each pond;
the volumetric BOD and COD loadings on anaerobic ponds;
the surface BOD and COD loadings on facultative ponds; and
the percentage removals of BOD, COD, suspended solids, nitrogen,
phosphorus, E coli and nematode eggs in each pond and in each series of
Monitoring and Evaluation of WSP 187
A simple first-order kinetic analysis may be undertaken if desired (Chapter 5).
The responsible local or central governmental agency should record and store
all the information on, and all the data collected from, each pond complex,
together with an adequate description of precisely how they were obtained, in
such a way that design engineers and research workers can have ready and
meaningful access to them.
Wastewater Storage and Treatment
Wastewater storage and treatment reservoirs (WSTR), also called effluent
storage reservoirs, were developed in Israel to enable the whole year’s treated
wastewater to be used for crop irrigation during the irrigation season (Juanicó
and Shelef, 1991, 1994; Juanicó and Dor, 1999; see also Barbagallo et al,
2003*; Friedler et al, 2003*). With an irrigation season of four months in
Israel, this means that three times the land area can be irrigated, and three
times the quantity of crops produced. In water-short areas, such as Israel and
also many parts of the developing world, this is a major agricultural advantage.
The use of WSTR thus maximizes the potential of wastewater re-use for crop
production (Chapter 21).
Single WSTR (Figures 16.1 and 16.2a) are mostly used in Israel as the main
crop is cotton – that is, the practice is restricted irrigation (Chapter 21) and
the principal microbiological quality for the treated wastewater is ≤1 or ≤0.1
human intestinal nematode eggs/l (Chapter 4). Israeli practice with single
WSTR is to treat the wastewater first in an anaerobic pond (Chapter 10). The
WSTR have a depth of 5–25 m and a volume equal to the volume of effluent
produced in the 8-month non-irrigation season, so that it is full immediately
before the irrigation season. During the irrigation season, the WSTR contents
are pumped out to the fields to be irrigated, while at the same time there is still
a continuous inflow of anaerobic pond effluent. As the irrigation season
progresses, there is a progressive deterioration in the quality of the WSTR
contents as the retention time in the reservoir of the anaerobic pond effluent
becomes correspondingly shorter (Liran et al, 1994). For restricted irrigation,
certainly for the irrigation of cotton, this does not matter, but for unrestricted
irrigation it would be a serious problem.
Wastewater Storage and Treatment Reservoirs 189
Figure 16.1 Single WSTR in Israel
If the local farmers wish to practise unrestricted irrigation – that is including
the irrigation of salad crops and vegetables eaten uncooked – then single
WSTR are inappropriate as the treated wastewater must contain no more than
1000 E coli/100 ml throughout the irrigation season (as well as meeting the
nematode egg guideline level) (Chapters 4 and 21). ‘Sequential batch-fed
WSTR’ are easily able to achieve this E coli requirement (Mara and Pearson,
1992; Juanicó, 1996*).
Depending on the length of the irrigation season, three or four sequential
batch-fed WSTR are used to store and treat the anaerobic pond effluent (Figure
16.2b). The WSTR are in parallel, and each is operated on a cycle of
‘fill–rest–use’. As soon as the contents of one reservoir are used, another is
brought into irrigation service and the one just emptied is refilled in readiness
for the next irrigation season (Table 16.1). E coli numbers decline rapidly
during the rest phase: in studies in northeast Brazil Pearson et al (1996c*)
found that in WSTR which received organic loadings of 126–162 kg BOD/ha
day during the fill phase, E coli numbers dropped from 106–107/100 ml of
anaerobic pond effluent to <1000/100 ml during the first 14 days of the rest
phase at 25°C.
Sequential batch-fed WSTR are an integral component of Shelef and Azov’s
(2000*) vision of high-efficiency pond systems for the 21st century. Figure
16.3 shows the sequential batch-fed WSTR system for the town of Arad
(population 22,000) in the Negev desert, Israel.
190 Domestic Wastewater Treatment in Developing Countries
Restricted Unrestricted
Note: (a) single WSTR for restricted irrigation; (b) sequential batch-fed WSTR for unrestricted
irrigation; and (c) hybrid WSP–WSTR system for both restricted and unrestricted irrigation. A,
anaerobic pond; F, facultative pond.
Figure 16.2 Wastewater Storage and Treatment Reservoir Systems
Given that farmers often want treated wastewater for both restricted and
unrestricted irrigation, a hybrid pond–reservoir system is an appropriate
alternative to sequential batch-fed WSTR which only produce treated
wastewater for unrestricted irrigation – that is, it is of too high a quality for
restricted irrigation. In the hybrid system (Figure 16.1c) the wastewater is
treated in anaerobic and facultative ponds. During the non-irrigation season,
the facultative pond effluent is used to fill a single WSTR. During the irrigation
season the facultative pond effluent is used for restricted irrigation, and the
WSTR contents for unrestricted irrigation (Mara and Pearson, 1999*).
If the number of nematode eggs in the facultative pond exceeds 1/l (or 0.1/l
when children under 15 years are exposed), then an additional short-retentiontime maturation pond would be required (Chapters 11 and 12).
Design a wastewater storage and treatment reservoir system for the wastewater
detailed at the end of Chapter 10. Assume the irrigation season is 6 months.
Wastewater Storage and Treatment Reservoirs 191
Table 16.1 Operational Strategy for Three Sequential Batch-fed WSTR for an
Irrigation Season of Six Months
Fill (1)b
Fill (1)
Fill (1/2)
Fill (1/2)
Fill (1/3)
Fill (1/3)
Fill (1/2)
Fill (1/2)
Fill (1/2)
Fill (1/2)
Fill (1/3)
Fill (1/3)
Fill (1/2)
Fill (1/2)
Fill (1)
Fill (1)
Fill (1)
Fill (1)
Fill (1/3)
Fill (1/3)
a July and August are the hottest months, so WSTR No 3 has the minimum rest period of two
months at this time. The other two WSTR have rest periods of four months to ensure E. coli die-off
to <1000 per 100 ml during the cooler months
b Proportion of monthly flow discharged into each WSTR
c WSTR volume expressed as multiple of monthly wastewater flow
Restricted irrigation
Pretreat the wastewater in the anaerobic pond calculated in Chapter 10.
Choose a single WSTR (Figure 16.2a), which must be full at the start of the
irrigation season and empty at the end of it. Thus the WSTR volume is equal
to 6 months wastewater flow:
V = (365/2) x 10,000
= 1,825,000 m3
Assuming a depth of 10 m, the WSTR mid-depth area is 18.25 ha.
Unrestricted irrigation
Pretreat the wastewater in the same anaerobic pond, and choose three
sequential batch-fed WSTR in parallel as shown in Figure 16.2b. The volume
of each WSTR is equal to four months wastewater flow (Table 16.1):
V = (365/3) x 10,000
= 1,216,700 m3
192 Domestic Wastewater Treatment in Developing Countries
Note: (a) view of Arad town in the Negev desert, (b) covered anaerobic pond, (c) WSTR in rest
phase, and (d) WSTR in use phase.
Figure 16.3 Sequential Batch-fed WSTR at Arad, Israel
Wastewater Storage and Treatment Reservoirs 193
Assuming a depth of 10 m, the mid-depth area of each WSTR is 12.17 ha. The
total mid-depth area, including that of the anaerobic pond, is 37 ha.
Restricted and unrestricted irrigation
Assume that the local farmers wish to use half the treated wastewater for
restricted irrigation and half for unrestricted irrigation. Choose the hybrid
WSP–WSTR system shown in Figure 16.2c. This comprises the anaerobic and
secondary facultative ponds calculated in Chapters 10 and 11, and the single
WSTR calculated above. The total mid-depth area is 22 ha.
Constructed Wetlands
Many natural wetlands have received domestic wastewater pollution for a long
time, and their productivity is increased as they have a large capacity for
absorbing domestic wastewater. Constructed wetlands are the engineer-made
equivalent of natural wetlands, and they are designed to reproduce and
intensify the wastewater treatment processes that occur in natural wetlands.
Thus rooted aquatic plants, often termed ‘macrophytes’, are grown either in
soil or, now more commonly, gravel beds which receive domestic wastewater
after primary treatment in, for example, an anaerobic pond. Constructed
wetlands are also called ‘reedbeds’ after the aquatic macrophyte most
commonly grown in them – Phragmites australis, the common reed. Other
plants commonly used include Schoenoplectus lacustris (bulrush), Typha
latifolia (cattail) and Juncus effusus (soft rush). Constructed wetlands are long,
narrow, shallow (ie almost plug flow) reactors (Figure 17.1) in which the
partially treated wastewater is treated further by natural wetland processes.
Fully comprehensive reviews of constructed wetlands for wastewater treatment
are given by IWA Specialist Group (2000) and Sundaravadivel and
Vigneswaran (2001*).
In industrialized countries, constructed wetlands are sometimes used to
create wildlife habitats (often with public access and visitor facilities) following
their use for wastewater treatment (eg Environmental Protection Agency,
1993*). There is nothing inherently wrong in this, of course, but it should be
realized that this is not wastewater treatment and therefore should be funded
separately – unless the wastewater treatment authority has the necessary funds
and wishes to derive some public-relations benefit from so doing. Wildlife
habitat creation is really an example of treated wastewater re-use.
In this chapter only gravel-based ‘subsurface-flow’ wetlands are discussed,
principally because of the major risk of mosquito (especially Coquillittidia spp
and Mansonia spp) breeding in ‘surface-flow’ wetlands in developing
There are two types of subsurface-flow wetlands: horizontal-flow and verticalflow units. Horizontal-flow wetlands are the most common; vertical-flow
Constructed Wetlands 195
Source: Courtesy of Dr Rebecca Stott
Figure 17.1 A 100-m Long Subsurface-flow Constructed Wetland in Egypt
wetlands are occasionally used in temperate climates as a secondary stage to
horizontal-flow wetlands. Only horizontal-flow wetlands are considered here.
Horizontal-flow wetlands
BOD, suspended solids and nutrients (N and P) are removed in subsurfaceflow wetlands by a combination of the mechanisms occurring in rock filters
treating effluents from waste stabilization ponds (Chapter 12) and mechanisms
due to the plants. Tanner (2001*) found no improvement in BOD, suspended
solids, phosphorus and faecal bacterial removals in planted wetlands compared
with unplanted control systems (which functioned therefore as rock/gravel
filters). There was, however, better nitrogen removal in the planted beds than
in the unplanted beds, but this was not primarily due to uptake by the plants,
but more to the accumulation of organic nitrogen in the bed sediments and to
the release of oxygen by the plants around their roots which is then used for
ammonia removal by nitrification, with removal of the nitrate so formed by
denitrification in the bulk anoxic zone of the gravel bed. Similar results were
reported by Ayaz and Akça (2001*) and Regmi et al (2003).
Horizontal subsurface-flow constructed wetlands are normally designed
for BOD removal, which is modelled by equation 5.10 for plug flow reactors,
but with the retention time given by:
196 Domestic Wastewater Treatment in Developing Countries
θ = Vcw/Q = εAcwDcw/Q
where ε is the porosity of the planted gravel bed (typically 0.4 for 25-mm
gravel), and Q is the mean flow, m3/day, given by equation 11.6 with e (net
evaporation) being taken here as the net evapotranspiration, mm/d. Thus
equation 5.10 becomes:
Le = Liexp{– k1[2εAcwDcw/(2Qi – 0.001eAcw)]}
The value of k1 depends not only on temperature, but also on the porosity as
bacterial growth occurs on the surface of the gravel in the gravel bed and the
amount of bacterial growth present changes with porosity, and also with
temperature, as follows (Reed et al, 1988; Corea, 2001*):
k1 = 68.6ε4.172(1.06)T–20
Corea (2001*) used equations 17.2 and 17.3 to determine the size of on-site
wetland systems for hotels in Sri Lanka treating the effluent from septic tanks
and anaerobic filters. For Li = 100 mg/l, Le = 10 mg/l, ε = 0.4, Dcw = 0.6 m
(the value most commonly used), T = 27°C, e = 0 and Q = 0.15 m3/day (the
wastewater flow per person), he determined the wetland area to be 0.64 m2
per person (this is close to the value of 0.66 m2 per person found in Egypt by
Williams et al, 1995). In the case of housing estates with only septic tanks,
Corea determined for Li = 150 mg/l and Le = 30 mg/l (the Sri Lankan effluent
BOD standard for surface water discharge) a wetland area of 0.45 m2 per
Reeds were used in the wetland systems for housing estates by Corea
(2001*), but for the hotel systems a variety of ornamental plants was used
(Figure 17.2); broad-leaved plants were used to reduce the effluent flow as
their evapotranspiration is high, but if the effluent was required to water the
hotel gardens, other plants were used and the effluent was treated further in a
small vertical-flow wetland prior to its use. Corea (2001*) should be consulted
for further details of these on-site hotel systems.
Suspended solids removal
Suspended solids are removed in subsurface-flow constructed wetlands by
entrapment in the gravel interstices and sedimentation. Reed and Brown
(1995) give the following equation for SS removal:
(SS)e = (SS)i[0.106 + 0.11(AHLR)]
Constructed Wetlands 197
Source: Corea, 2001*
Figure 17.2 A Horizontal-flow Constructed Wetland at a Hotel in Kandy,
Sri Lanka, Planted with Ornamentals
198 Domestic Wastewater Treatment in Developing Countries
where (SS)e and (SS)i are the effluent and influent suspended solids
concentrations, respectively, mg/l; and AHLR is the areal hydraulic loading rate
(= Qi/Acw), m/day.
Ammonia removal
Ammonia is removed in subsurface-flow constructed wetlands by several
factors, such as nitrification (and subsequent denitrification), plant uptake and
accumulation of organic nitrogen in the bed sediments. Huang et al (2000*)
give the following equation for ammonia-N removal in constructed wetlands:
Ce = Ciexp[ – 0.126(1.008)T–20θ]
where Ce and Ci are the effluent and influent ammonia concentrations,
respectively, mg N/l; and θ is given by equation 17.1. Equation 17.5 was
derived from constructed wetlands in the US, planted with Schoenoplectus and
Typha and treating septic tank effluent; bed temperatures were 6°C in winter
and 20°C in summer.
Helminth egg removal
Helminth egg removal in the gravel bed of horizontal-flow wetlands is very
efficient: Stott et al (1999*) found that all eggs were removed in a 100-m long
reedbed in Egypt, with most being removed in the first 25 m.
Horizontal-flow constructed wetlands are secondary treatment units – that is,
they must be preceded by a septic tank or anaerobic pond. Therefore their area
should be compared with that of a secondary facultative pond (Chapter 11).
Wetland area
• Li
Population served
= 90 mg/l (the BOD of the effluent of the anaerobic
pond detailed in Chapter 10),
= 18 mg/l (the same as the filtered BOD of the
effluent of the secondary facultative pond
detailed in Chapter 11),
= 25°C,
= 5 mm/day,
= 0.4,
= 0.6 m,
= 10, 000 m3/day, and
= 100, 000
Using equations 17.2 and 17.3 with the design parameter values used in
Chapter 11, the area of the wetland is calculated as follows, assuming
evapotranspiration = evaporation:
Constructed Wetlands 199
from equation 17.3:
k1 = 68.6 (0.4)4.172(1.06)5 = 2.0 day–1
rearranging equation 17.2:
Acw = [–2Qiln(Le/Li)]/[2k1ε Dcw – (0.001eln(Le/Li))]
= [– 2 x 10000 ln(18/90)]/[(2 x 2 x 0.4 x 0.6) – (0.001 x 5 ln(18/90))]
= 33 250 m2 = 0.33 m2 per person
Secondary facultative pond
The area of the secondary facultative pond detailed in Chapter 11 is 26,500
m2 – that is 0.27 m2 per person.
The horizontal-flow constructed wetland thus requires 22 per cent more
area than the secondary facultative pond; it also requires ~53 t of gravel. In
general, therefore, ponds are preferable to wetlands. Calculations of the kind
done above are necessary to demonstrate the preference for WSP over
constructed wetlands in developing countries, so that preferences to the
contrary (eg Juwarkar et al, 1995*; Haberl, 1999*; Kivaisi, 2001*) can be
logically rebutted. An exception to this is, for example, the small on-site hotel
systems with ornamental plants described by Corea (2001*).
Upflow Anaerobic Sludge
Blanket Reactors
Upflow anaerobic sludge blanket reactors (UASBs) are high-rate anaerobic
wastewater treatment units. They were developed in the 1970s by Professor
Gatze Lettinga at the University of Wageningen in The Netherlands, and they
have been extensively tested at full-scale in tropical and subtropical regions,
particularly in Brazil, Columbia and India (van Haandel and Lettinga, 1994;
Foresti, 2002*; for an excellent on-line introduction to UASBs, see Field and
Sierra, 2002*). They are used for the primary treatment of domestic
wastewaters and high-strength biodegradable industrial and agro-industrial
wastewaters. They have also been found satisfactory for the treatment of
mixed domestic and industrial wastewaters in Mauritius (Dean and Horan,
1995*). However, an alternative view of their performance and applicability
in developing countries (at least in Kanpur, India) is given by Sharma (2002*).
UASBs are reinforced-concrete structures (Figure 18.1), with a short
hydraulic retention time, of the order of 6–12 hours. As shown in Figure 18.2,
the raw wastewater, after screening and grit removal (Chapter 8), is distributed
as evenly as possible across the base of the reactor. It then flows upwards
through the sludge layer (termed the sludge ‘blanket’); this ensures intimate
contact between the wastewater and the anaerobic bacteria in the sludge
blanket, so aiding the anaerobic biochemical reactions detailed in Chapter 3
and thus increasing the efficiency of BOD removal in the reactor. The
wastewater, together with some active sludge particles, then rises through the
reactor, and, during this time, further BOD reduction occurs. The
wastewater–sludge suspension then reaches the ‘phase separator’, which is the
important characteristic of this type of anaerobic reactor: it divides the reactor
into its two constituent zones – the lower digestion zone and the upper settling
zone. As the wastewater–sludge suspension rises through the settling zone, its
upflow velocity decreases since, due to the outwardly inclined surface of the
phase separator, the flow area increases and the suspended sludge particles
settle out, mainly on to the inclined sides of the phase separator. Eventually
the weight of the accumulated sludge particles exceeds the frictional force
Upflow Anaerobic Sludge Blanket Reactors 201
Note: Top, general view; bottom, view showing top of UASB, including effluent overflow weirs.
Figure 18.1 An UASB at Ginebra, Valle del Cauca, Southwest Colombia,
Treating 10 l/s of Domestic Wastewater
202 Domestic Wastewater Treatment in Developing Countries
separator element
Anaerobic sludge blanket
Source: van Haandel and Lettinga (1994)
Figure 18.2 Schematic Diagram of an UASB
keeping them on the inclined surfaces, and they settle down to the sludge layer.
The phase separator thus maintains a high concentration of sludge particles in
the lower zone of the reactor, and enables an effluent with a very low
suspended solids concentration to be discharged from the reactor.
Biogas bubbles are collected under the phase separator, from where the
gas is easily extracted either for use (to generate electricity, for example) or for
flaring-off. Deflectors are placed between the phase separator units to prevent
any biogas bubbles entering the settling zone where they would hinder
UASBs are very efficient, as shown by the following equation for COD
removal (van Haandel and Lettinga, 1994):
R = 100(1 – θ–0.68)
where R is the percentage COD removal (which essentially equals the
percentage BOD removal for domestic wastewaters); and θ is the mean
hydraulic retention time, hours. Thus for θ = 6 h (the value recommended by
van Haandel and Lettinga for domestic wastewaters in the single-compartment
UASBs considered here), R = 70 per cent.
UASB design is described in full by van Haandel and Lettinga (1994), which
should be consulted for greater detail than the brief outline given here.
The upflow velocity (Vup, m/h) of wastewater in UASBs, which should not
exceed 1 m/h, is given by:
Upflow Anaerobic Sludge Blanket Reactors 203
Vup = Q/A = QD/AD = D/θ
where Q is the wastewater flow, m3/h; A is the surface area of the reactor, m2;
and D is the depth of wastewater in the reactor, m.
Thus, for Vup = 1 m/h and for θ = 6 h, D = 6 m. In practice the range of D
used is 4–6 m, so Vup is ≤1 m/h.
UASB reactor size is normally limited to 1000 m3. An UASB of this volume
with a retention time of 6 h can treat 4000 m3 of wastewater per day; for larger
flows additional UASB reactors are provided in parallel. UASBs are normally
rectangular in plan with a length-to-breadth ratio less than 4 to 1. Parallel
units share longitudinal walls.
Influent distribution
The raw wastewater, after screening and grit removal (Chapter 8), has to be
distributed as evenly as possible over the reactor base. Influent outlet devices
are located on the reactor base at the rate of one per 3–4 m2 of reactor base
area. Each of these devices has to receive an equal proportion of the raw
wastewater flow; this is ensured by having an influent distribution channel
which is closed at the far end and which has a number of 90º V-notches along
one side, with the number of notches being equal to the number of outlet
devices on the reactor base. The distribution channel is designed for a liquid
level in the channel 125 mm higher than the apex of the V-notch. Figure 18.3
shows the general arrangement of the influent distribution channel and
distribution boxes.
(a) circular inlet device
(b) rectangular inlet device
Influent distributor channel
To inlet points
To inlet
Section A–A’
Source: Adapted from van Haandel and Lettinga (1994)
Figure 18.3 Influent Distribution Channel and Distribution Boxes
204 Domestic Wastewater Treatment in Developing Countries
Phase separators
The phase separators divide the reactor into its two zones, the settling zone
and the sludge zone, and they separate the three phases occurring in the reactor
– gas, liquid and solid. Often the separators are submerged (Figure 18.4) as
this keeps the biogas under a small positive pressure. The sides of the
separators are inclined at 45–60° to the horizontal, and there should be an
overlap of at least 100 mm between the lower edge of the separator and the
deflector in order to prevent any escape of biogas into the settling zone. The
volume of the settling zone (ie the liquid volume outside the separators and
above their base) is 15–20 per cent of the total reactor volume.
The total base width of the separator units is around two-thirds of the
total width of the reactor (Figure 18.4). Thus the upflow velocity of the rising
wastewater–sludge suspension increases to 3Vup at the separator base level as
it passes between adjacent separator units, and then decreases to Vup at the
discharge level.
Effluent collection
The treated effluent is collected in a horizontal gutter which has regular Vnotches (two notches per m2 of reactor surface area for Vup = 1 m/h). A scum
guard is installed to prevent any floating solids leaving the reactor. The general
arrangement is the same as that used on primary and secondary sedimentation
tanks in conventional wastewater treatment plants (see Metcalf and Eddy, Inc,
Sludge production and drying
UASBs produce quite large amounts of waste sludge, ~0.2 kg/kg of BOD
removed. This is much less than is produced in conventional activated sludge
plants (~0.8 kg/kg), but much more than in anaerobic ponds (Chapter 10);
and UASBs produce waste sludge which needs to be disposed of continuously
(see Cavalcanti et al, 1999*), whereas anaerobic ponds only produce waste
sludge intermittently.
In warm climates, UASB waste sludge can be simply dewatered on drying
beds. Drying bed design is covered in detail by van Haandel and Lettinga
(1994); in essence the drying bed area required at temperatures ≥20°C is in the
range 0.01–0.015 m2 per person. The total drying bed area is divided into 3–5
individual beds; these can be covered with a simple roof to aid drying in the
rainy season. Sludge is wasted once every 1–2 weeks; it dries quite quickly and
without odour release, usually within 3–4 weeks. As it will contain some viable
Ascaris eggs, the dried sludge should be stored for at least three months before
it is applied to agricultural land. Alternatively, it can be landfilled immediately.
Upflow Anaerobic Sludge Blanket Reactors 205
(a) Submerged separator
Pgas = Patm + Ph
Vdi = V1
Settling zone
Vm = 0.31V1
Separator element
Separator element
(b) Separator with gas under atmospheric pressure
Separator element
Separator element
(c) Hybrid separator with opening for maintenance
Separator element
Separator element
Source: van Haandel and Lettinga (1994)
Figure 18.4 Details of a Submerged Phase Separator
206 Domestic Wastewater Treatment in Developing Countries
A 6-h UASB achieves a 70 per cent removal of BOD, but this is also achieved
by a 1-day anaerobic pond at 25°C (Chapter 10). Which is ‘better’? (This
question was asked, and answered, by Peña Varón et al, 2000*.) The UASB is
clearly smaller: it has only one quarter of the volume of the anaerobic pond.
However, it costs more to construct a 6-h UASB in reinforced concrete (even
reinforced brickwork) than it costs to construct (essentially excavate) a 1-day
anaerobic pond. Furthermore, the saving in land area is insignificant when the
area of the secondary facultative pond needed to treat the anaerobic effluent
(from either the UASB or the anaerobic pond) and the area of the drying beds
for the UASB sludge are taken into account. This can be illustrated as follows:
UASB and facultative pond
For a wastewater flow of 10,000 m3/day, a 4-m deep 6-h UASB has an area of
625 m2; the drying beds for the UASB sludge have an area of (100,000 persons
x0.15 m2/person) = 1500 m2; and the secondary facultative pond designed in
Chapter 11 has an area of 26,500 m2 – that is a total mid-depth area of 28,000
Anaerobic and facultative ponds
For the same wastewater flow, a 3-m deep 1-day anaerobic pond has an area
of 3340 m2 giving, with the same facultative pond, a total mid-depth area of
29,840 m2 – only 6.6 per cent more than the UASB–facultative pond
combination. If maturation ponds are used, then the land saving due to using
UASBs, rather than anaerobic ponds, is even less.
Thus, in almost all situations anaerobic ponds, and in particular high-rate
anaerobic ponds (Chapter 10), will be preferred to UASBs. An exception to
this will be when high-strength industrial or agro-industrial wastewaters are to
be treated, or pretreated prior to discharge to sewer; in such cases the finance
for construction and the skilled personnel for operation and maintenance are
more readily available – as, for example, at the potato chip factory in Egypt
reported by El-Gohary and Nasr (1999*; see also Field and Sierra, 2002*).
UASBs are also useful to treat wastewaters with a high sulphate concentration
as the resulting hydrogen sulphide in the biogas can be readily removed in a
chemical scrubber or a ‘bioscrubber’ (Nishimura and Yoda, 1997*).
Biofilters (also called trickling filters, percolating filters and bacteria beds) are
an old process for the secondary treatment of domestic wastewater dating from
the beginning of the 20th century (Institution of Water and Environmental
Management, 1988), and there are many thousands of biofilters in use in both
industrialized and developing countries. However, for use in warm climates
(whether in industrialized or developing countries) they should now only be
used with modern fly control techniques (see below). Biofilters produce high
quality effluents (eg <20 mg BOD/l and <30 mg SS/l) without requiring large
areas of land or consuming vast quantities of electricity. In many situations in
developing countries they are much more appropriate than activated sludge.
Biofilters comprise a 2–3 m deep bed of 50–100 mm rock (Figure 19.1).
The settled wastewater – that is, the effluent from either a primary
sedimentation tank, or in developing countries more appropriately an
anaerobic pond (Broome et al, 2003*) – is distributed mechanically over the
rock medium (Figure 19.2) and it percolates down through the medium to be
collected in an underdrain system at the base of the bed. A microbial film
develops on the surface of the rock and the bacteria in this ‘biofilm’ oxidize
the wastewater organics (ie remove the BOD) as the settled wastewater trickles
down through the bed.
The biofilm develops in thickness as the bacteria in it grow on the settled
wastewater. Eventually it becomes too thick and some is sheared off (often
described as ‘sloughed off’) by the wastewater flow. The solids (sometimes
called ‘humus solids’ or simply ‘humus’) have to be removed in a secondary
sedimentation tank (a ‘humus tank’) or in a sedimentation or maturation pond,
so that the final effluent has a sufficiently low suspended solids concentration.
Biofilters are plug flow reactors in which BOD removal is reasonably well
represented by first-order kinetics (Chapter 5). The effluent BOD (after
sedimentation) is thus given by equation 5.10:
208 Domestic Wastewater Treatment in Developing Countries
Note: A, inlet pipe; B, underdrain blocks; C, effluent channel; D, outlet pipe.
Figure 19.1 Sectional Perspective View of a Circular Biofilter Showing the
Rotating Wastewater Distributor and Filter Medium
Le = Lie–k1θ
The retention time is a function of the specific surface area S of the filter
medium (its surface area per unit volume, m2/m3), the gross filter volume V
(m3) and the flow rate Q (m3/day), and is equal to distance/velocity. The
distance travelled is a slightly zigzag path which is proportional to the depth
(D, m) of the biofilter – let this be αD (where α >1). To calculate the velocity
(= flow/unit area), consider a 1-m cube of the filter medium:
the flow applied to the cube is Q/A, where A is the cross-sectional area of
the biofilter (m2), and
the cross-sectional area available for effective flow is the area immediately
adjacent to the film of the microbial growth. In the 1-m cube there are S
m2 of surface area. Since the depth of the cube is 1 m, its effective linear
cross-sectional dimension (ie ‘flow perimeter’) is S m. If the flow thickness
is d m, the cross-sectional area used for the flow is xSd, where x is <1 and
is a combined measure of the non-availability and non-use of the
theoretical flow area (Sd) for actual flow (non-availability arises from the
presence of the microbial film and non-use is due to the non-ideal
distribution of the wastewater over the whole cross-section).
The flow velocity is flow/area – that is (Q/A)/(xSd), and the retention time θ is
distance/velocity, so that:
(Q/A) / xSd
Biofiltration 209
Figure 19.2 Distribution of Settled Wastewater on to a Rectangular Biofilter
Equation 5.10 therefore becomes:
Le = Liexp(– KVS/Q)
where K is a ‘modified rate constant’ ( = k1αxd), m/day; and V is the biofilter
volume ( = AD), m3.
Since k1 is temperature-dependent (equation 5.8), K must also vary with
temperature. Studies at pilot-scale in England (Water Pollution Research
Board, 1973) showed that:
KT = 0.037(1.08)T–15
The proportion of voids (‘porosity’) in a filter comprising 50–100 mm
aggregate is ~40 per cent of the gross volume. When a strong wastewater is
applied to such a filter, excessive biofilm growth occurs and this can lead to
blockage of the filter and ‘ponding’ of the wastewater on the surface of the
filter. Experience has shown that, if a final effluent of BOD ≤20 mg/l and SS
≤30 mg/l is required, the BOD loading that can be safely be applied to a singlepass filter is ≤15 kg/m3day. However, if some of the clarified effluent is
returned to the filter inlet, 2–3 times the normal loading for a single-pass filter
can be applied with the production of a satisfactory effluent.
Recirculation serves not only to dilute the settled wastewater but, more
importantly, to reduce the rate of biofilm growth and to increase the hydraulic
210 Domestic Wastewater Treatment in Developing Countries
stripping of the film. It also ensures that more of the available surface area is
used for waste oxidation by providing a more uniform hydraulic distribution
across the filter and also a more uniform vertical distribution of the microbial
film. The recirculation ratio (ie the ratio of the recycled flow to the settled
wastewater flow) is generally in the range 0.5–6, with 1 being most commonly
Alternative design equations
There are several empirical or semi-empirical design equations for biofilters,
mainly derived from biofilter plants in the US (see Baker and Graves, 1968;
Horan, 1990; Metcalf and Eddy, Inc, 1991). One of the most commonly used
is the National Research Council (1946) equation; its original form, in US
customary units, is:
1 + 0.0561(W/VF)0.5
E20 =
where E20 is the BOD removal efficiency at 20°C; W is the BOD load applied
to the filter, lb/day; V is the filter volume, 1000 ft3; and F is the corrected
recirculation factor (dimensionless) given by:
F =
(1 + 0.1R)2
where R is the ratio Qr/Qi (Qr is the recirculation flow rate and Qi is the
influent wastewater flow rate prior to recirculation). If there is no
recirculation, then F = 1.
The efficiency at T°C is given by Metcalf and Eddy, Inc (1991) as:
ET = E20(1.035)T–20
In metric units, equation 19.4 becomes:
E20 =
1 + 0.225(W/VF)0.5
where W is now in g/day and V in m3 (1 lb = 454 g and 1000 ft3 = 28.32 m3).
Note that W/V = LiQi/V (where Li is the influent BOD before recirculation,
mg/l; and Qi is in m3/day) = λv, g/m3 day (equation 10.1).
Equations 19.5–19.7 are used for design.
Biofiltration 211
Bacteria are not the only organisms to live in biofilters. There is a whole
complex ecosystem of protozoa, metazoa, worms – and fly larvae. Unless
controlled, swarms of newly emergent adult ‘filter flies’ (such as Psychoda and
Sylvicola) and chironomid midges can make life unpleasant for nearby
residents (Surrey Advertiser, 1998*).
Fly control is best achieved by covering the rock medium with high-density
polyethylene netting (Palmhive Technical Textiles, 2002*) (Figure 19.3), which
prevents gravid female flies from entering the biofilter to lay their eggs – no
eggs, no larvae, no flies (Burridge, 2000). Of course, a few flies will get in and
lay their eggs; the resulting newly emergent flies are best attracted to, and killed
by, electronic fly killers (P & L Systems, 2002*).
The humus (suspended solids) in the biofilter effluent must be removed prior to
discharge. This is achieved in secondary sedimentation tanks (for the design of
these, see Metcalf and Eddy, Inc, 1991 and Horan, 1990) or in sedimentation
ponds (Chapter 20). If the biofilter effluent is to be used for restricted crop
irrigation, a sedimentation pond is used (rather than a secondary sedimentation
tank) in order to achieve ≤1 nematode egg/l or ≤0.1 egg/l if children under 15
years are exposed. If it is to be used for unrestricted irrigation, a series of
maturation ponds (Chapter 12) is necessary to reduce the E coli count to
≤1000/100 ml; the first maturation pond should be 0.5 m deeper than the others
to allow for settlement, digestion and storage of the humus solids.
Figure 19.3 Rectangular Biofilters Covered with High-density Polyethylene
Netting to Control Fly Nuisance
212 Domestic Wastewater Treatment in Developing Countries
Design a biofilter to treat the effluent of the anaerobic pond designed in
Chapter 10.
At 25°C KT is given by equation 19.3 as:
K25 = 0.037(1.08)25–15
= 0.08 m/day
Rearrangement of equation 19.2 gives:
V = – [Q ln(Le/Li)][KTS]–1
For Le = 30 mg/l and S = 40 m2/m3 (a typical value for 50–100 mm rock)
V = – [10 000 x ln(30/90)][0.08 x 40]–1
= 3500 m3
Choose two biofilters in parallel, each 36 m in diameter and 1.8 m deep.
Simple Activated Sludge Variants
Aerated lagoons are activated sludge units operated without sludge return.
Historically they were developed from waste stabilization ponds in the northern
US, where mechanical aeration was used to supplement the algal oxygen supply
in winter. It was found, however, that soon after the aerators were put into
operation the algae disappeared and the microbial community quickly came to
resemble that of activated sludge. Aerated lagoons (Figure 20.1), especially
those operating at short retention times in warm climates, are designed as
completely mixed non-return activated sludge units. Floating aerators (Figure
20.2) are used to supply the necessary oxygen and mixing power.
Aerated lagoons can treat either raw wastewater (after preliminary
treatment – Chapter 8), or settled wastewater, for example, anaerobic pond
effluent (Chapter 10), as at Melbourne, Australia (Chapter 9), Ben Slimane,
Morocco (Kouraa et al, 2002*) and Daqahla, Egypt (El Sharkawi et al,
1995*). BOD removals above 90 per cent are achieved at short retention times
(2–6 days); retention times less than 2 days are not recommended as they are
too short to permit the development of a healthy flocculent sludge (even so the
activated sludge concentration is only 200–400 mg/l, in contrast to the
2000–6000 mg/l found in conventional systems and oxidation ditches). In
common with all activated sludge systems, aerated lagoons are not particularly
effective in removing faecal bacteria: reductions are only 90–95 per cent and
further treatment may therefore be necessary (Chapters 12, 21 and 22).
The rate of bio-oxidation of wastewater in an aerated lagoon has been found
to be approximated well by first-order kinetics – equation 5.7:
Le =
1 + k1θ
In the case of aerated lagoons, the BOD of the effluent (Le) is due to two
separate fractions: (1) the small amount of the influent waste not oxidized in
214 Domestic Wastewater Treatment in Developing Countries
Source: Courtesy of Aeration Industries International, Inc.
Figure 20.1 An Aerated Lagoon
the lagoon, and (2) the bacterial cells synthesized during oxidation (see Figure
3.4). These fractions are generally referred to as the ‘soluble’ and ‘insoluble’
BOD, respectively. It is convenient (and conceptually more correct) to apply
first-order kinetics only to the removal of the soluble fraction:
Se =
1 + κθ
where Se is the soluble BOD in the effluent (ie the fraction of the influent BOD
which escapes oxidation), mg/l; and κ is the first-order rate constant for soluble
BOD removal, day–1.
All the influent BOD is assumed to be soluble (ie Si = Li). The retention
time in the lagoon is 2–6 days, with 4 days the most usual value. A typical
design value for κ is 2.5 day–1 at 20°C; its values at other temperatures can be
estimated from the equation (Metcalf and Eddy, Inc, 1991):
κT = 2.5(1.06)T–20
The quantity of bacteria synthesized in the lagoon is related to the quantity of
soluble BOD removed:
Simple Activated Sludge Variants 215
Source: Courtesy of Aeration Industries International, Inc.
Figure 20.2 Floating ‘Aire-O2 Triton’ Aerator-mixer with Two Propellers –
One for Mixing and one for Fine Bubble Distribution of the Air Which is
Supplied by a Compressor
where X is the cell concentration in the lagoon, mg/l; and Y is the yield
coefficient (dimensionless, and defined by this equation as the mass of cells
formed per unit mass of soluble BOD consumed).
Y is typically 0.6–0.7. On a finite time basis, say one retention time,
equation 20.3 can be rewritten for the whole lagoon as:
Y (Li – Se)V
where V is the lagoon volume, m3.
The rate of cell synthesis must be balanced by the sum of the rates at which
the cells leave the lagoon in the effluent and at which they die in the lagoon.
The rate at which the cells leave the lagoon is QX where Q is the flow through
the lagoon. The rate at which some of the cells in the lagoon die is proportional
to the quantity of cells present; it is usually given as bXV where b is the rate of
autolysis in day–1 (equation 3.4; typically b = 0.07 day–1 at 20°C). Thus:
216 Domestic Wastewater Treatment in Developing Countries
Y (Li – Se)V
= bXV + QX
Rearranging and writing V/Q as θ:
Y (Li – Se)
1 + bθ
This quantity of cells X can be converted to an equivalent ultimate BOD by
considering the chemical equation for their complete oxidation:
C5 H7 NO2 + 5O2 5CO2 + 2H2O + NH3
Thus 1 g of cells has an ultimate BOD of (5 x 32/113) = 1.42 g. Since
BOD5/BODu ≈ 2/3, 1 g of cells has a BOD5 of 0.94 g. Therefore the total
effluent BOD is given by:
Le = Se + 0.94X
Oxygen requirement
The quantity of oxygen required for bio-oxidation (RO2, kg/h) is the amount
of total (ie soluble + insoluble) ultimate BOD removed:
RO2 = [1.5(Li – Le)Q] x 10–3/24
Substituting equation 20.7 (rewritten in terms of ultimate BOD):
RO2 = [1.5(Li – Se)Q – 1.42XQ] x 10–3/24
that is, the oxygen requirement is the ultimate soluble BOD removed less the
ultimate BOD due to the cells wasted in the effluent.
Nitrification (the autotrophic bacterial conversion of ammonia to nitrite and
then to nitrate – Chapter 3) may occur in aerated lagoons (and almost certainly
in oxidation ditches – see later in this chapter) in warm climates. Nitrification
requires 3.1 mg O2 per mg ammonia-N nitrified (Chapter 3); it also requires
~7 mg of alkalinity as CaCO3 per mg ammonia-N nitrified.
Simple Activated Sludge Variants 217
The kinetic equations are as follows:
θs = 1.6(µT – b)–1
where θs is the solids retention time (the ‘sludge age’) in the aerated lagoon,
days – since there is no solids (sludge) recirculation in an aerated lagoon θs =
θ, the V/Q hydraulic retention time, days; µT is the specific growth rate of the
nitrifying bacteria at T°C, day–1 (given by equation 20.11); and b is their
endogenous decay rate, day–1 (taken as 0.2 day–1). The coefficient 1.6 in
equation 20.10 is the ratio of the peak to average influent ammonia-N loads
(in, for example, kg N/day) – that is, it is a factor of safety to allow the
nitrifying bacterial population to become accustomed to fluctuations in
influent ammonia-N concentrations.
µT = 0.8(1.086)T–20
s(N) + Ce
)( K
s(DO) + DO
where Ce is the required effluent ammonia-N concentration, mg/l; Ks(N) is the
Monod half-saturation constant for ammonia-N (see equations 3.4 and 20.12),
mg/l; DO is the in-lagoon dissolved oxygen concentration, mg/l (typically 1–2
mg/l); and Ks(DO) is the Monod half-saturation constant for DO, mg/l (typically
0.5 mg/l).
The value of Ks(N) varies with temperature, as follows (Horan, 1990):
Ks(N)(T) = 0.733(1.1246)T–20
A minimum hydraulic retention time (= sludge age in aerated lagoons) of ~3
days is needed at temperatures >20°C to permit nitrification to occur.
The oxygen required for nitrification is given by:
RO2 = 3.1(Ci – Ce)Q x 10–3/24
where RO2 is the nitrification oxygen requirement, kg/h; and Ci and Ce are the
influent and effluent ammonia-N concentrations, respectively, mg/l.
The total oxygen requirement is the sum of the requirements for biooxidation (equation 20.8 or 20.9) and for nitrification (equation 20.13).
Aerator performance
The aerators must supply both sufficient oxygen for bio-oxidation and
sufficient power to mix the lagoon contents. The power required for mixing
218 Domestic Wastewater Treatment in Developing Countries
(Pm, W/m3) is ~3 W/m3 for the ‘Triton’ aerator-mixer shown in Figure 20.2
(Aeration Industries International Inc, 2002*). For conventional surface
aerators (of the type common in conventional activated sludge units), Pm is
much higher; Horan (1990) gives the equation:
Pm = 5 + 0.004X
where X is the cell (ie suspended solids) concentration in the lagoon, mg/l. As
X is typically in the range 200–400 mg/l, the power required for mixing is
~6 W/m3 for conventional surface aerators (ie around twice that used by
modern aerator-mixers).
Usually more power is required for mixing than for oxygen supply for biooxidation, but this must be checked as follows: manufacturers certify that their
aeration equipment has an ‘oxygen transfer rate’ of so many kg O2/kWh under
standard test conditions – these are: tap water as the test liquid at 20°C and
initially with zero dissolved oxygen concentration. This standard rating has to
be corrected for conditions in the aerated lagoon, as follows (Nogaj, 1972):
OL = O0[α][(1.024)T–20]
[ βCC
s(T,A) –
where OL is the oxygen transfer rate in the aerated lagoon, and O0 is the
oxygen transfer rate under standard test conditions.
The first correction term is to allow for the nature of the liquid to be
α = the ratio of the oxygen transfer rate in the wastewater to that in
tap water at the same temperature (typically for domestic
wastewaters, a = 0.8).
The second term is an Arrhenius temperature correction term. The third term
is a dissolved oxygen correction term which allows for the difference between
the DO concentration in the lagoon (typically 1–2 mg/l) and that adopted in
the standard rating test (initially zero):
= O2 saturation concentration (ie O2 solubility) in distilled
water at temperature T and altitude A. The values of Cs(T)
at sea level (760 mm Hg) are given in Table 20.1 for
T = 15–30°C. The correction for altitude is made by
considering the mean air pressure PA (mm Hg) at that
Simple Activated Sludge Variants 219
Table 20.1 Solubility of Oxygen in Distilled Water at Sea Level at
Various Temperatures
Temperature (ºC)
Solubility (mg/l)
Source: Montgomery et al (1964)
= Cs(T,0) (PA/760)
= O2 saturation concentration in distilled water at 20°C and
at sea level (= 9.08 mg/l),
= DO concentration in the lagoon (1–2 mg/l), and
= the ratio of the O2 saturation concentration in the
wastewater to that in distilled water (typically for domestic
wastewaters, β = 0.95).
Modern aerators are rated at ~1.5–2 kg O2/kWh.
The construction of aerated lagoons is essentially the same as that for waste
stabilization ponds (Chapter 13). The major differences are: greater depths
(usually 3–5 m), steeper embankment slopes (1 to 1.5–2) and the provision of
a high-density liner to prevent scour by the turbulence induced by the aerators.
Effluent treatment
Consider a typical domestic waste with Li = 300 mg/l. For θ = 4 d and T =
20°C and assuming κ = 2.5 day–1, b = 0.07 day–1 and Y = 0.65, equations 20.1,
20.5 and 20.7 give:
220 Domestic Wastewater Treatment in Developing Countries
Se =
= 27 mg/l
1 + κθ
1 + (2.5 x 4)
0.65 (300 – 27)
Y (Li – Se)
= 140 mg/l
1 + bθ
1 + (0.07 x 4)
Le = Se + 0.94X = 27 + (0.94 x 140) = 160 mg/l
Thus the BOD of the lagoon effluent is 160 mg/l, but >80 per cent of this is
due to the bacterial cells present. If these cells (or most of them) are removed
from the effluent prior to discharge, the effluent BOD will be considerably
reduced. There are two ways to do this: (1) discharge into a series of
maturation ponds, or (2) discharge into a sedimentation pond.
Discharge into maturation ponds
If the final effluent is to be re-used in agriculture or aquaculture (Chapters 21
and 22), then discharge into a series of maturation ponds is the favoured
option in order to achieve the required microbiological quality. The maturation
ponds are designed as described in Chapter 12, with the assumption either that
the E coli count is reduced in the aerated lagoon by 90 per cent, or – in the
case of restricted irrigation – that there is no helminth egg reduction in the
aerated lagoon.
As the aerated lagoon effluent is high in suspended solids (~200–400 mg/l),
provision has to be made in the first maturation pond for the settlement and
subsequent digestion of these solids. This is simply done by increasing the
depth in this pond to 2 m. Regular desludging will be required (see Chapter
Discharge into a sedimentation pond
This is the preferred option when the final effluent is discharged into a surface
water. The sedimentation pond should be designed according to the following
requirements (Metcalf and Eddy, Inc, 1991):
the hydraulic retention time must be sufficient to allow the settleable solids
to settle,
sufficient volume must be provided for sludge storage,
algal growth must be minimized, and
odours must be controlled.
Hydraulic retention time. Two considerations are relevant here: first, the
minimum retention time for solids sedimentation is 6 h; and second, to
minimize algal growth, the maximum retention time is 2 days. Generally a
retention time of 1 day is used.
Odour control. To control any odours emanating from the anaerobic
digestion of the settled solids, there must be a minimum liquid depth of 1 m
Simple Activated Sludge Variants 221
above the sludge layer at its maximum depth (ie just before the pond is
desludged). Often a depth of 1.5 m is used.
On the basis of the above two considerations, the liquid volume and
surface area of the sedimentation pond can be calculated, as follows:
Volume (V, m3):
V = Qθ
Surface area (A, m2)
A = Qθ/DL
where Q is the inflow, m3/day; θ is the hydraulic retention time, typically 1
day; and DL is the liquid depth required for odour control, typically 1.5 m.
Sludge storage. The calculation of the volume required for the storage of
digested sludge is done in the following three steps:
1 Calculate the total mass of suspended solids added per year:
M = 365Q(SSi – SSe) x 10–3
where M is the total mass of suspended solids added, kg/year; Q is the inflow,
m3/day; SSi and SSe are the influent and required effluent suspended and solids
concentrations, respectively, mg/l (= g/m3); and the factor 10–3 converts g to
The value of SSi in equation 20.19 is computed as follows:
SSi = SSRW + X
where SSRW is the suspended solids concentration in the raw wastewater, mg/l;
and X is the cell concentration given by equation 20.6 (this assumes that the
SSRW are not biodegraded in the lagoon).
2 Calculate the mass of fixed and volatile suspended solids added, MFS and
MVS (kg/year), assuming that 70 per cent of the total solids are volatile:
MFS = 0.3M
MVS = 0.7M
222 Domestic Wastewater Treatment in Developing Countries
3 Calculate the volume required for sludge storage:
(i) assuming that within a period of one year the volatile solids are reduced
by 75 per cent, the following equation gives the mass of volatile solids
present at the end of year n:
(Mvs)n = 0.25nMvs
thus the mass of total solids present after n years is:
(M)n = nMFS + 0.25nMvs
(iii) determine the pond depth for sludge storage (DS, m), assuming that the
sludge compacts to an average value of 15 per cent solids and that the
density of the accumulated sludge is 1060 kg/m3:
Ds = [(M)n /A]/[0.15 x 1060]
Thus the sedimentation pond has a surface area of A m2 (equation 20.18) and
a total depth of (DL + DS). The term n in equations 20.23 and 20.24 is the
required desludging frequency – once every n years.
Aerated lagoons or anaerobic ponds?
Often in developing countries when aerated lagoons have been installed (not
uncommonly for the ‘reason’ that they are more ‘sophisticated’ and therefore
‘better’ than WSP), the wastewater treatment authority finds that it cannot
afford to pay the electricity bill for them, so the aerators are switched off –
with the result that the 4-day lagoons quickly convert themselves into 4-day
anaerobic ponds and BOD removal remains high, especially at temperatures of
≥20°C (Tables 10.1 and 10.2). (The aerators are generally only switched on
again when important or foreign visitors are expected.) The question to be
asked, therefore, is: shouldn’t anaerobic ponds have been used in the first
place? Or, when anaerobic ponds are followed by aerated lagoons: shouldn’t
anaerobic and secondary facultative ponds have been used?
The answer to the first question is straightforward: if there is no effluent
ammonia-N requirement (and so nitrification in an aerated lagoon is not a
consideration), an honest comparison between the two technologies (especially
when high-rate anaerobic ponds are considered) will always favour anaerobic
ponds – particularly when annual electricity costs are properly quantified (and
transparently presented) at the project prefeasibility stage (remember that
money spent on electricity is money gone forever, but money spent on land is
an investment – Chapters 6 and 9). The answer to the second question also
requires an honest comparison to be made between the two systems: an aerated
Simple Activated Sludge Variants 223
lagoon requires less land than a secondary facultative pond [the retention
times, especially at temperatures >20°C, may not be too different, but the
difference in depth (3–4 m for aerated lagoons vs 1.5 m for facultative ponds)
is significant], but it requires a large annual expenditure on electricity. Could
this expenditure be used not for electricity, but to finance a loan to buy more
land or to convey the wastewater further out of town to an area where land is
less expensive?
Design example
Design an aerated lagoon and sedimentation pond system to treat a unit flow
of 1000 m3/day of domestic wastewater which has BOD and SS concentrations
of 300 and 400 mg/l, respectively. Effluent requirements are ≤20 mg BOD/l
and ≤30 mg SS/l. The design temperature is 20°C.
(a) Aerated lagoon
Take θ = 4 days, κ = 2.5 day–1, b = 0.07 day–1 and Y = 0.65. From equations
20.1, 20.6 and 20.7:
Se =
= 27 mg/l
1 + κθ
1 + (2.5 x 4)
0.65 (300 – 27)
Y (Li – Se)
= 140 mg/l
1 + bθ
1 + (0.07 x 4)
Le = Se + 0.94X = 27 + (0.94 x 140) = 160 mg/l
Lagoon size: assume a depth of 3 m, so the lagoon mid-depth area is given by:
A = Qθ/D = 1000 4/3 = 1340 m2
Choose two lagoons in parallel, each 26 m square at mid-depth.
Aeration: estimate the quantity of oxygen required from equation 20.8:
RO2 = 1.5(Li – Le)Q = 1.5(300 – 160)1000 = 210,000 g/day = 8.8 kg/h
Assume that the aerators have a standard rating of 1.8 kg O2/kWh and correct
for field conditions from equation 20.15 (assume α = 0.8, β = 0.95, cL = 1 mg/l
and zero altitude):
OL = O0α(1.024)T–20
( βCC
s(T,A) –
224 Domestic Wastewater Treatment in Developing Countries
= 1.8 x 0.8 x 1 x
x 9.08) – 1
( (0.95 9.08
= 1.2 kg O2/kWh
The aerator power required for bio-oxidation = 8.8/1.2 = 7.3 kW.
Choose a ‘Triton’ aerator-mixer which uses 3 W/m3 for complete mixing.
As the lagoon volume is (1340 x 3) m3, the power required for mixing is
therefore 12 kW. Thus the power requirement for mixing is greater than that
for oxygen supply.
(b) Sedimentation ponds
Pond area: take θ = 1 day and DL = 1.5 m. From equation 20.18:
A = Qθ/DL = 1000 x 4/1.5 = 2670 m2
Sludge storage: from equation 20.20 the concentration of suspended solids
entering the sedimentation pond is:
SSi = SSRW + X = 400 + 140 = 540 mg/l
From equations 20.19, 20.21 and 20.22:
M = 365Q(SSi – SSe) x 10–3
= 365 x 1000 (540 – 30) x 10–3
= 186,000 kg/year
MFS = 0.3M = 0.3 x 186,000 = 56,400 kg/year
MVS = 0.7M = 0.7 x 186,000 = 131,600 kg/year
From equations 20.24 and 20.25 and choosing n = 3 years:
(M)n = nMFS + 0.25nMVS
= (3 x 56 400) + (0.25 x 3 x 131,600)
= 268,000 kg
Ds = [(M)n/A]/[0.15 x 1060]
= (268,000/2670)/(0.15 x 1060)
= 0.63 m
Simple Activated Sludge Variants 225
Thus the sedimentation pond area is 2670 m2 and its total depth is 1.5 + 0.63
= 2.13 m (say, 2.2 m).
Oxidation ditches are a direct modification of conventional activated sludge
(Baars, 1962; Barnes et al, 1983; Environmental Protection Agency, 2000b*).
Their essential operational features are that they receive raw wastewater (after
preliminary treatment) and provide longer retention times: the hydraulic
retention time is commonly 0.5–1.5 days and that for the solids 20–30 days.
The latter, achieved by recycling >95 per cent of the activated sludge, ensures
minimal excess sludge production and a high degree of mineralization in the
small amount of excess sludge that is produced. Sludge handling and treatment
is almost negligible since the small amounts of waste sludge can be readily
dewatered without odour on drying beds. The other major difference is in
reactor shape: the oxidation ditch is a long continuous channel, usually oval in
plan and 2–3 m deep (Figure 20.3). The ditch liquor is aerated by several
aerators (eg ‘Triton’ aerator-mixers – Figure 20.2), which impart a velocity to
the ditch contents of 0.3–0.4 m/s to keep the activated sludge in suspension.
The ditch effluent is discharged into a secondary sedimentation tank to permit
solids separation and sludge return and to produce a settled effluent with low
BOD and SS. Removals consistently >95 per cent are obtained for both BOD
and SS.
The concentration of total SS in the ditch is 3000–5000 mg/l. In order to
prevent the concentration exceeding this range, the return sludge flow is
diverted to the drying beds for a short period each day; this period is best
determined by operational experience (a simple field check on the ditch SS
concentration is to fill a 1000-ml graduated cylinder to the mark with the ditch
liquor; if the solids concentration is 3500–4500 mg/l, the volume of sludge
which settles in 30 minutes should be ~200 ml). Alternatively the sludge
wastage rate may be estimated by considering the solids retention time.
The oxidation ditch was developed in The Netherlands to provide small
communities of 200–15,000 people with wastewater treatment facilities at the
same cost per person as conventional activated sludge works serving much
larger populations. At present there are few oxidation ditches in developing
countries since waste stabilization ponds are usually more favourable, both in
terms of costs and faecal bacterial removal, although where there is a reliable
electricity supply but insufficient land for ponds they are being increasingly
used. As noted in Chapter 9, oxidation ditches were found by Arthur (1983*)
to be cheaper than aerated lagoons and biofilters, and also cheaper than waste
stabilization ponds when land costs exceeded US$50,000–150,000 per ha.
Oxidation ditch design is purely empirical. The depth is 2–3 m and the volume
is dependent on the retention time which in turn is based on the sludge loading
226 Domestic Wastewater Treatment in Developing Countries
Source: Courtesy of Aeration Industries International, Inc.
Figure 20.3 Typical Oxidation Ditch Installation
factor, γ. This is the mass of BOD applied to the ditch liquor suspended solids
per day; it is measured in g BOD per g solids per day (ie day–1). The mass of
BOD entering the ditch is LiQ g/d, where Li is the influent BOD (mg/l, = g/m3)
and Q the flow (m3/day); the mass of suspended solids in the ditch is SV where
S is the ditch liquor suspended solids concentration (mg/l) and V the ditch
volume (m3). Thus the sludge loading factor is given by:
Simple Activated Sludge Variants 227
or, since V/Q = θ:
Design values commonly used in Europe are γ = 0.05 day–1 and S = 4000 mg/l
which give, for a typical domestic waste (Li = 300 mg/l), a retention time of
1.5 day. However, in warm climates much higher loadings, and therefore
shorter retention times, are possible (Arceivala and Alagarsamy, 1970): a
comparison between the design criteria used in temperate and warm climates
is given in Table 20.2.
Oxygen is required at a rate of ~1.5 g O2/g BOD applied. Such a rate of supply
includes an allowance for the endogenous respiration of the sludge and
maintains aerobic conditions along the entire length of the ditch.
Sludge wastage
To control the rate of solids accumulation in the ditch, a proportion of the
ditch solids must be wasted each day. The rate of wastage is governed by the
desired solids retention time (SRT): thus, if the desired SRT is θs days, then
100/θs per cent of the ditch solids must be wasted each day. This can be
achieved by diverting 100/θs per cent of the flow from the ditch directly to
waste. However, it is more usual to waste the sludge from the sludge return
line from the secondary sedimentation tank. The percentage 100/θs must
therefore be corrected for the change in concentration due to sedimentation.
Noting that the inflow concentration to the sedimentation tank is the ditch SS
concentration S, the quantity of waste sludge to be diverted to the drying beds,
expressed as a percentage of the raw wastewater flow Q, is:
Table 20.2 Design Criteria for Oxidation Ditches in India and Europe
Sludge loading factor (day–1)
Aeration requirement (kg O2/kg BOD applied)
Excess sludge production (g/hd day)
Area of sludge drying beds (m2/hd)
Overall land requirement (m2/hd)
Source: Arceivala and Alagarsamy (1970)
228 Domestic Wastewater Treatment in Developing Countries
( SS )
where SR is the the solids concentration in the sludge return line (= underflow
concentration from the sedimentation tank).
For example, choosing θs = 30 days and S = 4000 mg/l and assuming that
SR = 50,000 mg/l, the sludge wastage rate would be 0.27 per cent. This
illustrates the extremely low rate of sludge production in oxidation ditches.
Since oxidation ditches are likely to be used in developing countries only for
large (and probably only very large) flows, they should be constructed with
vertical walls of reinforced concrete, and they should be as deep as possible: 3
m wherever feasible, but at least 2 m.
Design example
Design an oxidation ditch scheme to serve a population of 100,000. The
effluent flow is 80 l/person day and the BOD contribution is 40 g/person day.
The design temperature is 20°C and the ditch site is at sea level.
(a) Oxidation ditch
Flow Q = 80 x 100,000 l/day = 8 000 m3/day
Influent BOD Li = 40 x 1000/80 = 500 mg/l
From Table 20.2 choose γ = 0.20 day–1 and S = 4000 mg/l. Then from equation
500 x 8000
= 5000 m3
4000 x 0.20
θ = V/Q = 5000/8000 = 0.625 day = 15 h
The oxygen requirement is taken as 1.5 x the BOD load:
RO2 = (1.5 x 40 g/person day) x (100,000 people)
= 6000 kg/day = 250 kg/h
The aerator rating is now corrected for field conditions using equation 20.15:
Simple Activated Sludge Variants 229
OL = O0α(1.024)T–20
( βc c
s(T,A) –
= 1.8 x 0.8 x 1 x
x 9.08) – 1
( (0.95 9.08
= 1.2 kg O2/kWh
The power for the oxygen supply is therefore 250/1.2 = 208 kW. The power
for complete mixing is (5000 m3 x 3 W/m3) – that is 15 kW. Thus the required
power is that for oxygen supply.
(b) Sedimentation tank
The area is based on a peak overflow rate of 25 m3/m2 day and a retention
time of 2 h at peak flow. Since the sludge flow is returned to the ditch (ignoring
the small volume wasted), the peak overflow is simply taken as 2 x inflow for
a population of 100,000 (Chapter 8). Therefore:
area of tank
depth of tank
= (2 x 8000)/25 = 640 m2
= (2 x 8000) (2/24)/640 = 2 m
Choose two oxidation ditches in parallel, each followed by a sedimentation
tank 20 m in diameter and 2 m deep.
(c) Drying beds
Assume an area of 0.025 m2 per person (Table 20.2); thus:
Drying bed area
= 0.025 x 100 000 = 2500 m2.
Wastewater Re-use in Agriculture
Water is becoming scarcer and scarcer in developing countries and also in parts
of some industrialized countries. In arid and semi-arid areas especially, but in
fact everywhere, wastewater is simply too valuable to waste. It contains scarce
water and valuable plant nutrients, and crop yields are higher when crops are
irrigated with wastewater than with freshwater (Table 21.1). Farmers also save
on artificial fertilizers: in Mexico, for example, this saving is around
US$135/ha year, which is a significant amount of money for subsistence
farmers (Future Harvest, 2001*).
Treated wastewater is used for crop irrigation in many parts of the world.
Israel, for example, uses over 65 per cent of its wastewater in this way, and
plans to use over 90 per cent by 2010 (Friedler, 2001*); and in the desert areas
of the US, such as Arizona and California, there are large wastewater re-use
schemes (Asano, 1998). Australia is another good example (Dillon, 2000), and
India has been using wastewater for irrigation for nearly 100 years (Shende
et al, 1988). Mexico City, the second largest city in the world, uses all its
wastewater for irrigation (Duron, 1988).
As water becomes scarce, the competition between urban water demand
and agricultural water demand increases. In the not too distant future it will
have to be ‘water for cities and treated wastewater for agriculture’. This must
be recognized and planned for, and this means that wastewater has to be
considered as part of a country’s water resources; it is in fact a very dependable
Table 21.1 Crop Yields for Wastewater and Freshwater Irrigation
Irrigation water
Raw wastewater
Settled wastewater
Waste stabilization
pond effluent
Fresh water + NPK
Moong beans
Note: Yields in t/ha year
Source: Shende (1985)
Wastewater Re-use in Agriculture 231
water resource (Asano, 2002), especially during droughts (Bruins, 1999). In
his analysis of water resource availability, development and use in the Middle
East, Beaumont (2000*) found that each cubic metre of water used by industry
and the service sector generates at least 200 times more wealth than when used
by agriculture, and he concludes that ‘as water shortages increase, many
countries will be best served by the reallocation of irrigation water to meet the
growing water needs of the urban regions’. Irrigation water will still, of course,
be needed, and treated urban wastewater will become one of the principal
irrigation water sources near cities in developing countries: this is happening
now, but it will have to happen much more in the near future.
Guidelines for assessing the environmental impacts of proposed
wastewater re-use schemes are given (within the context of US and Californian
environmental law, but applicable in principle elsewhere) by Kontos and Asano
(1996*). An extremely useful document for government departments planning
wastewater use is the ‘Queensland Water Recycling Strategy’ (Environmental
Protection Agency, 2001*); Queensland is a tropical state of Australia whose
population is expected to increase by 40 per cent in the next 20 years, so
placing great pressures on existing water resources.
Of course, if treated (or even untreated) wastewater is discharged into a
river, it is eventually re-used downstream for irrigation and/or urban supply.
This is termed ‘indirect re-use’, but here we are concerned with ‘direct re-use’
– that is, the use of treated wastewater for crop irrigation, or for aquaculture
(Chapter 22), without prior discharge to river.
Wastewater re-use in Islamic countries
In 1978 the Leading Council of Islamic Scholars of Saudi Arabia issued a
‘fatwa’ (an authoritative religious ruling) on wastewater re-use. Prior to this it
was unclear whether treated wastewaters were Islamically ‘pure’ and there had
been a consequent reluctance to use treated wastewater for crop production
and landscape and urban green space irrigation (Abderrahman, 2001). As
noted by Farooq and Ansari (1983) there are three ways in which ‘impure’
water may be transformed into ‘pure’ water:
self-purification of the water (eg removal of the impurities by
addition of pure water in sufficient quantity to dilute the impurities, and
removal of the impurities by the passage of time or physical effects (eg
sunlight and wind).
The first and third of these transformations are essentially similar to those
achieved by modern wastewater treatment processes, especially waste
stabilization ponds.
232 Domestic Wastewater Treatment in Developing Countries
Crop irrigation with untreated wastewater, though often practised (and
generally liked by farmers as it is so ‘rich’ in nutrients and organics), cannot be
recommended as it causes too much disease in those who work in the
raw-wastewater-irrigated fields and in those who consume the rawwastewater-irrigated crops, especially salad crops and vegetables eaten
uncooked. The health risks are simply too high when raw domestic wastewater
is used to irrigate crops. (Irrigation with untreated wastewater is discussed in
more detail in the final section of this chapter.)
Actual and potential health risks
For an actual health risk to be caused by crop irrigation with wastewater, all
of the following four steps must occur:
an infective dose of an excreted pathogen reaches the field (or the pathogen
multiplies in the field to form an infective dose);
the infective dose reaches a human host;
the host becomes infected; and
the infection causes disease (or further transmission).
If only the first three of these occur and not the fourth, then the risk is only a
potential risk and not an actual risk. (These four steps also apply to
aquacultural re-use (Chapter 22), with the substitution of ‘aquaculture pond’
for ‘field’.)
Finally, but very importantly, the use of treated wastewater for crop
irrigation (or for aquaculture pond fertilization) is only of public health
importance if it causes a risk that is higher than that to which people are
already exposed – that is, if it results in an excess incidence or prevalence of
disease, or an excess intensity of infection, above what may be termed the
‘background’ level.
Restricted and unrestricted irrigation
Restricted irrigation refers to the irrigation of all crops except salad crops and
vegetables which may be eaten uncooked. Unrestricted irrigation includes the
irrigation of salad crops and vegetables eaten uncooked. This distinction is
important as the groups of people exposed are different: for restricted
irrigation it is necessary to protect the health of the fieldworkers, especially
children working or playing in wastewater-irrigated fields, and of neighbouring
communities (who may be at risk from wastewater aerosols from spray or
sprinkler irrigation systems); and for unrestricted irrigation the health of the
crop consumers must also be protected.
Wastewater Re-use in Agriculture 233
The epidemiological evidence
Shuval’s appraisal
Professor Hillel Shuval and his colleagues at the Hebrew University of
Jerusalem undertook an extensive review of the available credible
epidemiological evidence for actual health risks resulting, and not resulting,
from crop irrigation with wastewater (Shuval et al, 1986*). Their main
conclusions were:
irrigation with raw wastewater causes an excess prevalence of Ascaris and
hookworm infection, and an excess intensity of these infections, both in
those who work in raw-wastewater-irrigated fields (Figures 21.1 and 21.2)
and in those who consume raw-wastewater-irrigated crops eaten uncooked
(Figure 21.3);
irrigation with treated wastewater does not cause any excess prevalence of
Ascaris infection among crop consumers (Figure 21.4); and
irrigation with raw wastewater also causes excreted bacterial disease in
those who consume raw-wastewater-irrigated crops eaten uncooked – the
evidence for this is mainly the outbreak of cholera in Jerusalem in 1970
when raw-wastewater-irrigated lettuce was on sale; when this was removed
from the market, the epidemic ceased (Fattal et al, 1986).
Prevalence (%)
Source: Shuval et al (1986*).
Figure 21.1 Excess Prevalence of Ascaris and Hookworm Infections in
‘Sewage Farm’ Workers in India Compared with a Control Group (a ‘sewage
farm’ practises raw wastewater irrigation)
234 Domestic Wastewater Treatment in Developing Countries
Medium-to-heavy infections (%)
Source: Shuval et al (1986*).
Figure 21.2 Excess Intensity of Ascaris and Hookworm Infections in Sewage
Farm Workers in India Compared with a Control Group
The 1989 WHO guidelines
In 1989 the World Health Organization published guidelines for the
microbiological quality of treated wastewaters used for crop irrigation (and
also for aquaculture – Chapter 22) (World Health Organization, 1989*). Based
principally on the conclusions of the above epidemiological review by Shuval
et al (1986*), these guidelines were: for restricted irrigation, ≤1 human
intestinal nematode eggs/l; and for unrestricted irrigation, the same egg
guideline level and ≤1000 E coli/100 ml. They require some modification to
take into account: (a) the extensive new epidemiological research, principally
that by Dr Ursula Blumenthal at the London School of Hygiene and Tropical
Medicine and her colleagues in Mexico, and (b) recent applications of
quantitative microbial risk analysis to wastewater re-use (see later in this
chapter). (Revised guidelines will be published by WHO, 2004* – see below.)
Blumenthal’s appraisal
Blumenthal et al (2000*) reviewed all the credible epidemiological,
experimental and risk analysis evidence published since Shuval et al (1986*).
Their main conclusions, together with those from more recent studies
(Blumenthal et al, 2001; Blumenthal and Peasey, 2002), were:
Wastewater Re-use in Agriculture 235
Ascaris prevalence (%)
Note: Farmers on the outskirts of Eastern Jerusalem used raw wastewater for crop irrigation,
including salad crop irrigation. Prior to the creation of the State of Israel in 1948, these crops were
on sale in Western Jerusalem and Ascaris prevalence was high. Following partition of the city in
1948 these crops were no longer on sale in Western Jerusalem and Ascaris prevalence declined
dramatically. When the city was reunited in 1966 the crops were again available and Ascaris
prevalence rose once more. Following the outbreak of cholera in the city in 1970, the irrigation of
crops with raw wastewater was banned and Ascaris levels dropped again.
Source: Shuval et al (1986*).
Figure 21.3 Ascaris Prevalence among Residents of Western Jerusalem,
(a) Restricted irrigation
• E coli guideline: epidemiological studies in Israel, Mexico and the US
showed that faecal coliform levels of 106/100 ml of treated wastewater
were associated with excess viral infection (but not disease), but not
at faecal coliform levels of ≤104/100 ml – thus suggesting that ≤105
E coli/100 ml would be an appropriate guideline value.
• Nematode egg guideline: studies in Mexico also showed that an egg level
of 1 per litre of treated wastewater was insufficiently protective for
children under the age of 15 years – thus indicating that a stricter guideline
of ≤0.1 egg per litre was more appropriate when children under 15 are
exposed, through either working or playing in wastewater-irrigated fields.
(b) Unrestricted irrigation
• E coli guideline: epidemiological studies in Mexico showed that when the
faecal coliform level in treated wastewater was 3 x 104/100 ml, there was
an excess risk of disease of 6 x 10–3 per person per year, but quantitative
microbial risk analysis (see later in this chapter) shows that when the
E coli level is 1000/100 ml, the risk of infection is ~10–4 per person per
year – that is <10–3 per person per year, which is the WHO tolerable risk
of infection from drinking fully treated drinking water.
236 Domestic Wastewater Treatment in Developing Countries
Raw wastewater
Treated wastewater
Ascaris prevalence (%)
Note: In Darmstadt (Da), and especially its suburb Griesheim (Gr), raw wastewater was used for
crop irrigation and Ascaris prevalence was very high. In Berlin (B), where treated wastewater was
used for irrigation, Ascaris prevalence was no higher than in other cities which used freshwater for
irrigation (Marburg, M; Bad Homburg, BH; Wiesbaden, W; Dillenburg, Di; and Giesen, Gi).
Treatment in Berlin was conventional treatment with both primary and secondary sedimentation
which removed almost all the Ascaris eggs.
Source: Shuval et al (1986*).
Figure 21.4 Ascaris Prevalence among Residents of Selected German Cities
Immediately After the Second World War
Nematode egg guideline: studies in Mexico suggested that the egg level
should be reduced to ≤0.1 per litre, as for restricted irrigation, in order to
protect children under 15 when (as expected) their fieldworker-parents
bring home ‘unrestricted’ produce direct from the fields or when such
produce is bought in local shops (but not in shops in neighbouring towns
or cities).
Experimental wastewater-irrigation of salad crops
(a) Helminth eggs
Stott et al (1994*) report on two separate sets of experiments, in one of which
lettuces were irrigated with waste stabilization pond effluents containing
various levels of human intestinal nematode eggs per litre (Ascaris, Trichuris
and Necator); and in the other lettuces were irrigated with freshwater
containing various numbers of eggs of the chicken roundworm, Ascaridia galli,
and the lettuces were then fed to immunosuppressed chicks which were
Wastewater Re-use in Agriculture 237
monitored for Ascaridia infection (these chicken–Ascaridia experiments were
done as a model for human–Ascaris infection).
The first set of experiments was conducted in northeast Brazil. Lettuces
were spray-irrigated with raw wastewater (160–200 eggs/l) and the effluents
from an anaerobic pond (14–18 eggs/l), a primary facultative pond (0.2–0.4
egg/l) and a tertiary maturation pond (0 egg/l). After five weeks irrigation the
lettuces irrigated with raw wastewater and the anaerobic pond effluent had
counts of 30–60 eggs/plant and 0.6 egg/plant, respectively, but those irrigated
with the two effluents having <1 egg/l had no eggs on their leaves.
In the second set of experiments conducted in glasshouses in West
Yorkshire, England, lettuces were spray-irrigated with treated wastewater
seeded with ≤1, 10 and 50 unembryonated viable A galli eggs/l. After five
weeks irrigation the mean egg counts per individual lettuce plant were 0.04,
0.24 and 0.64, respectively, but 54–78 per cent of the eggs were inviable. When
lettuces with these levels of egg contamination were fed to immunosuppressed
chicks, at a rate of one lettuce plant per chick per week for six weeks, none
developed Ascaridia infection. Thus the risk to consumers of salad crops
contaminated with very low numbers of nematode eggs at, or shortly after,
harvest is likely to be minimal.
(b) Faecal bacteria
In a series of experiments over three consecutive years in southern Portugal,
radishes and lettuces were drip- and furrow-irrigated with waste stabilization
pond effluent (Bastos and Mara, 1995*). To simulate unfavourable conditions,
the use of these ‘clean’ irrigation techniques was counterbalanced by the choice
of salad crops grown (one is a root crop and the other grows close to the soil).
The pond effluent quality was 2–5 x 103 E coli per 100 ml – that is, a little
higher than the WHO guideline for unrestricted irrigation and this provided
the basis for field-testing the guideline. At harvest, E coli levels on the crops
were ~103 and ~104 per 100 g (fresh weight) for radishes and lettuces,
respectively (ie within the range of acceptability for ready-to-eat foods – see
Gilbert et al, 2000* and below). Salmonella numbers were very low in the
pond effluents and rarely detected on either crop. These results thus provide
support for the WHO guideline level of ≤1000 E coli per 100 ml for
unrestricted irrigation.
Protozoan pathogens
Melloul et al (2002*) found that children aged 2–14 living in El Azzouzia, the
raw-wastewater-irrigation area of Marrakech, Morocco had a higher prevalence
of giardiasis and amoebiasis than children of the same ages living in a
neighbouring area where freshwater was used for irrigation: 67 per cent (Giardia
intestinalis, 39 per cent; Entamoeba histolytica, 28 per cent), compared with
26 per cent (Giardia, 20 per cent; E hystoytlica, 6 per cent). Younger children
(2–8) and boys (2–14) had higher prevalences than older children (9–14) and
girls (2–14); and children in El Azzouzia whose fathers were farm workers had
higher prevalences than those whose fathers were not. There was also a higher
238 Domestic Wastewater Treatment in Developing Countries
incidence of Salmonella infections in the children from El Azzouzia than in those
from the other area (21 per cent, compared with 1.1 per cent). This study is the
first to provide good evidence for the transmission of protozoan infections where
raw wastewater is used for crop irrigation, and it reinforces recommendations
that only treated wastewater should be used for irrigation.
The new WHO guidelines
The World Health Organization’s new guidelines for the microbiological
quality of wastewaters used for crop irrigation will be published in 2004
(World Health Organization, 2004*). It is expected that the new
recommendations will be that:
only treated wastewater be used for crop irrigation;
for restricted irrigation, treated wastewaters contain ≤105 E coli per 100
ml, and ≤1 human intestinal nematode egg per litre or, if children under 15
years are exposed, ≤0.1 egg/l; and
for unrestricted irrigation, treated wastewaters contain ≤1000 E Coli per
100 ml, and ≤1 human intestinal nematode egg per litre or, if local children
under 15 years are exposed, ≤0.1 egg/l.
1000 E coli per 100 ml
The guideline value of ≤1000 E coli per 100 ml created more than a little
controversy when it was originally introduced in 1989 (see Shelef, 1991). It
has to be asked, therefore, if 1000 E coli per 100 ml is sufficiently protective
of consumers’ health.
This is an important question to ask because:
The State of California (1978*) requires ≤2.2 total coliforms per 100 ml
for unrestricted irrigation;
The World Health Organization (1973) required ≤100 E coli per 100 ml
in 80 per cent of samples for unrestricted irrigation; and
The US Environmental Protection Agency and the US Agency for
International Development require zero E coli per 100 ml for unrestricted
irrigation (and this is a recommendation not only for the US, but for
developing countries as well) (Environmental Protection Agency, 1992 –
these guidelines are currently under revision).
So is 1000 E coli per 100 ml acceptable? The answer to this question is Yes,
because (Mara, 1995):
In the US, river water can be used for unrestricted irrigation if it contains
≤1000 E coli per 100 ml (Environmental Protection Agency, 1973).
In the European Union bathing (ie whole body immersion) is permitted if
recreational waters contain ≤2000 E coli per 100 ml (Council of the
European Communities, 1976 – this Directive is currently being revised).
Wastewater Re-use in Agriculture 239
The International Commission on Microbiological Specifications for Foods
(1974) allows foods eaten uncooked to contain up to 100,000 E coli per
100 g fresh weight, although preferably <1000.
In the European Union, foods are permitted to contain high E coli and
coliform numbers, for example, shellfish eaten raw (such as oysters) can
contain up to 230 E coli per 100 g live weight, and dairy products are
allowed to have high total coliform numbers: milk, 500/100 ml; butter,
1000/100 g; ice cream, 10,000/100 g; and soft cheese, 10,000,000/100 g;
and ‘hard cheese made from raw milk’ can contain up to 10,000,000
E coli/100 g (Council of the European Communities, 1991b, 1992).
In the United Kingdom ready-to-eat foods (sandwiches, salads, pizzas,
desserts, etc) can contain up to 10,000 E coli/100 g (Gilbert et al, 2000*).
Table 12.2 shows that when faecal coliform numbers are reduced in a
series of waste stabilization ponds to 7000/100 ml, there are no
salmonellae and no campylobacters, and only very low numbers of
rotaviruses and other enteroviruses.
Vaz da Costa Vargas and Mara (1988) found that local lettuce on sale in a
market in Portugal contained around 1,000,000 faecal coliforms per
100 g, and the faecal coliform numbers on lettuce irrigated with
conventionally treated wastewater containing 6 x 106 faecal coliforms per
100 ml fell to around 2 x 104 per 100 ml (ie better than the above ICMSF
requirement) five days after irrigation ceased.
Quantitative microbial risk analysis (discussed at the end of this chapter)
shows that the health risks associated with irrigation with treated
wastewater containing 1000 E coli per 100 ml are lower than the tolerable
risk of infection of 10–3 per person per year from drinking fully treated
drinking water (World Health Organization, 2003*).
So the WHO guideline level of no more than 1000 E coli per 100 ml is
perfectly satisfactory. It has been accepted legally in France, for example
(Conseil Supérieur d’Hygiène Publique, 1991; see also Faby et al, 1999*).
Water engineers, who are used to the requirement of zero E coli per 100
ml of treated drinking water, may instinctively think that 1000 E coli per 100
ml of irrigation water is simply too high. Yet as shown above, foods in
industrialized countries are allowed to have higher numbers of coliforms and
E coli. It is always worthwhile determining E coli numbers in or on foods,
especially those eaten uncooked, which are on sale in local markets – water
engineers (and indeed the local people) will be very surprised at the high levels
commonly found.
Options for public health protection
The best option for public health protection is to treat wastewaters to the
WHO guideline values given above, and this is most appropriately done in
waste stabilization ponds (Chapters 9–12), wastewater storage and treatment
reservoirs (Chapter 16), or other treatment processes (Chapters 17–20)
240 Domestic Wastewater Treatment in Developing Countries
Co W
ns ork
op Fiel
um er
po d/
w as
ex ate tecr r/
No protective
Crop restriction
and human
treatment and
exposure control
treatment and
crop restriction
Key to level of contamination (outer bands)/Risk (inner bands)
Pathogen flow
I Pond treatment II Conventional treatment
Source: Blumenthal et al (1989).
Figure 21.5 Generalized Model Showing the Levels of Relative Risk to
Human Health Associated with Different Combinations of Control Methods
for the Use of Wastewater in Agriculture and Aquaculture
supplemented with maturation ponds (Chapter 12). However, when ‘full’
treatment to the WHO guideline values is not possible for some reason (usually
financial), then there are still ways in which public health can be protected.
These are:
partial treatment,
crop restriction,
the method of wastewater irrigation, and
human exposure control.
These can be combined in various ways to protect the fieldworkers and/or the
crop consumers, as shown in Figure 21.5.
Wastewater Re-use in Agriculture 241
Partial treatment and crop restriction
These two go together as partial treatment here means treatment to achieve ≤1
or ≤0.1 egg/l and ≤105 E coli/100 ml – that is, for restricted irrigation. In fact
this will often be the preferred level of treatment and re-use purpose as partial
treatment obviously costs less than full treatment (ie to ≤1000 E coli/100 ml)
(see the design example in Chapter 12). However, this must be agreed with the
local farmers who have to undertake not to use the partially treated
wastewater to irrigate salad crops and vegetables eaten uncooked (ie for
unrestricted irrigation). Farmer participation in the whole wastewater re-use
project cycle – planning, design, implementation and management – is essential
for the success of re-use projects (see van Vuren, 1998). Engineers (and
planners) clearly need to talk to farmers.
Method of wastewater application
The most common methods of wastewater application in developing countries
are flood irrigation and furrow irrigation. Another method is ‘localized’
irrigation, a term which includes drip irrigation (also called trickle irrigation)
and ‘bubbler’ irrigation. Drip irrigation is extremely economical with irrigation
water: the wastewater is applied in plastic pipes fitted with emitter buttons
from which the wastewater is intermittently discharged to wet only the root
zone of the plants (Figure 21.6), so evaporative losses from the soil surface are
minimized. It is a relatively expensive technology and can normally be justified
only when irrigation water, including treated wastewater, is very scarce and
therefore has a very high value. However, low-cost drip irrigation systems are
being developed, at a cost of around one quarter of conventional systems,
especially for small-scale farmers in developing countries (Polak et al, 1997*;
see also Intermediate Technology Consultants, 2003*).
Drip irrigation with treated wastewater has the advantage over the other
methods of wastewater application in that the World Health Organization
(1989*, 2004*) does not specify any microbiological quality requirements for
the treated wastewater when applied to the field in this way. There is a
problem, however, with emitter clogging. This is commonly caused by soil
algae (which are attracted to the emitter by the nutrients in the wastewater
effluent), rather than the algae in waste stabilization pond effluents.
Commercially available emitters vary in their clogging potential, and a careful
choice of emitter has to be made when irrigating with treated wastewater
(Taylor et al, 1995*).
Bubbler irrigation is especially useful for the wastewater irrigation of trees,
such as fruit trees and nut trees. Instead of emitters, short vertical pipes are
used; the height of these pipes decreases along the length of the distribution
pipe to compensate for frictional head losses, so that each tree receives the
same amount of treated wastewater (Hillel, 1987*).
Human exposure control
Fieldworkers (and their employers) need to be aware of the health risks
associated with treated wastewater re-use. In particular, they should wear
242 Domestic Wastewater Treatment in Developing Countries
Figure 21.6 Drip Irrigation of Cotton with Maturation Pond Effluent at
Nicosia, Cyprus
rubber boots to protect themselves from hookworm infection, and they should
have hand-washing facilities available to them as they leave the field. This is
very important, although few employers provide either boots or hand-washing
facilities, and few fieldworkers will wear boots (since it is easier to work
barefoot in wet soils). However, they should be encouraged to do so, as a
healthy workforce is a productive workforce.
Consumers can protect themselves by using hygienic food preparation
practices: by cooking wastewater-irrigated vegetables and rigorously
washing wastewater-irrigated crops eaten uncooked (salad crops, fruits and
certain vegetables). Fruit and vegetable growers should also decontaminate
their produce; advice is given by the World Health Organization (Beuchat,
Irrigation with wastewater can, at least potentially, damage crops. Compliance
with the Food and Agriculture Organization’s recommendations for the quality
of waters used for irrigation (Ayers and Westcot, 1989*; see also Westcot,
1997*) will effectively eliminate such risks. For domestic wastewaters the most
relevant FAO recommendations are for the following five parameters, and
these need to be regularly checked during the irrigation season:
Wastewater Re-use in Agriculture 243
electrical conductivity
sodium absorption ratio
boron concentration
total nitrogen concentration
Electrical conductivity
This is used as a measure of the dissolved salts concentration, which represents
the ‘salinity hazard’. Wastewaters contain more salts than drinking waters as
we consume table salt (sodium chloride) far in excess of our physiological
requirements. For safe irrigation (Figure 21.7) the electrical conductivity
should be less than 75 mS/m at 20°C (the unit of conductivity is millisiemens
per metre, though it is often expressed as decisiemens per cm, but this is not
an SI preferred unit; it is necessary to specify the temperature of measurement
as conductivity varies greatly with temperature).
Irrigation with wastewater that is too saline causes interference with the
capacity of a plant’s roots to absorb water and nutrients, and therefore reduces
crop yields. Nevertheless, saline wastewaters can be used for irrigation even
though the crop yield may be reduced (it is better to produce some crops than
none), and certain crops commonly irrigated with treated wastewaters (for
example, cotton) are tolerant of medium-to-high levels of salinity. For further
details, the FAO review by Rhoades et al (1992*) should be consulted.
Sodium absorption ratio
This is used as a measure of the ‘sodium hazard’ of an irrigation water. The
SAR, which is dimensionless, is defined as:
[0.5([Ca+] + [Mg2+])]0.5
where [Ca+], [Mg2+] and [Na+] are the concentrations of calcium, magnesium
and sodium ions in the irrigation water, expressed in milliequivalents per litre.
Concentrations of these ions in mg/l are converted to meq/l by multiplying by
0.050, 0.082 and 0.044 for calcium, magnesium and sodium, respectively.
For safe irrigation, the SAR should be less than 18 (Figure 21.7). Irrigation
with waters with higher SAR values causes ‘sodium saturation’ of the soil:
calcium and magnesium atoms in the clay minerals which make up the soil are
displaced by sodium ions in the irrigation water, especially treated wastewater
with its higher sodium levels than the local drinking water. Eventually the soil
becomes saturated with sodium, and a sodium-saturated soil is difficult to
work (when dry it forms hard unmanageable clods), its internal drainage is
seriously affected, and the crops have difficulty in absorbing nutrients and
244 Domestic Wastewater Treatment in Developing Countries
3 4 5 6 7 8 9 10
3 4 5
Sodium adsorption ratio
Sodium alkali hazard
Conductivity (mS/m (mS/m at 25°C)
Salinity hazard
Note: Waters in regions A and B are acceptable for almost all irrigation purposes; those in regions
C should be avoided wherever possible; and those in the shaded area should not be used at all.
Source: Adapted from United States Department of Agriculture (1954).
Figure 21.7 Classification of Irrigation Waters Based on Conductivity and
Sodium Absorption Ratio
This element is present in wastewater from perborates in domestic detergents.
Most crops are tolerant of 2 mg B/l, but citrus fruit trees and deciduous nut
trees can only tolerate ~0.5 mg B/l (Table 21.2).
Total nitrogen
Too much nitrogen can reduce crop yields or cause crop damage. There may
be a luxuriant growth of the non-useful parts of the crop (eg large green leaves
on maize plants), or certain crops (eg lettuce) can suffer from ‘leaf burn’ –that
is the edges of their leaves turn brown. Most crops can tolerate 30 mg total
N/l, but some only 5 mg total N/l. Crop tolerances to total nitrogen are given
in the FAO guidelines.
Wastewater Re-use in Agriculture 245
Table 21.2 Recommended Maximum Concentrations of Boron in Irrigation
Waters According to Crop Tolerance
Boron concentration (mg B/l)
Lemon, blackberry
Avocado, grapefruit, orange, apricot, peach,
cherry, plum, fig, grape, walnut, pecan, cowpea,
Garlic, sweet potato, wheat, barley, sunflower,
beans, strawberry, peanut
Sweet pepper, pea, carrot, radish, potato,
Lettuce, cabbage, celery, turnip, oats, maize,
Sorghum, tomato, alfalfa, parsley, beetroot,
sugar beet, cotton
Source: Ayers and Westcot (1989*)
The permissible pH range for irrigation waters is 6.5–8.4, which does not
present a problem for treated domestic wastewaters.
Industrial effluents
If municipal wastewaters with a high proportion of industrial effluents are to
be re-used in agriculture, then a more detailed physicochemical analysis is
necessary as crop health is affected by many more parameters (eg heavy metals
– Table 21.3). The FAO guidelines should be consulted for further details
(Ayres and Westcot, 1989*; Westcot, 1997*). However, heavy metal
accumulation in crops irrigated with domestic wastewater in India has been
found to be lower than permissible levels, despite the wastewater having been
used for irrigation at the same site for ~30 years (Yadav et al, 2002*).
Waste stabilization ponds (Chapters 9–12) and wastewater storage and
treatment reservoirs (Chapter 16) are two excellent treatment options prior to
wastewater re-use in agriculture. They can easily achieve the required
microbiological quality and, when treating domestic wastewater, also achieve
the required physicochemical qualities (Table 21.4). In the soil, the pond algae
act as ‘slow release’ fertilizers and so contribute to increased crop yields and
soil organic matter, thus improving the water-holding capacity of the soil.
Other treatment processes usually require additional treatment to achieve
the required microbiological quality in, for example, maturation ponds
246 Domestic Wastewater Treatment in Developing Countries
Table 21.3 Recommended Maximum Metal Concentrations in Irrigation
Maximum concentration (mg/l)
a Toxic to rice at 0.05 mg/l
b Toxic to citrus fruits at 0.075 mg/l
Source: Ayers and Westcot (1989*)
(Chapter 12). There are other disinfection techniques, but these are not
recommended as:
in the case of ultra-violet disinfection, the high-intensity ultra-violet lamps
are expensive imported items;
in the case of chemical disinfection, the chemicals are expensive (eg ozone)
or needed in high doses (eg up to 20 mg chlorine/l), and even then they
will not be effective in reducing the numbers of helminth eggs and
protozoan cysts;
faecal bacterial regrowth may occur: a few resistant bacteria may survive
to multiply in an environment that is much less competitive as most of the
other faecal and non-faecal bacterial will have been killed; and
environmentally harmful compounds can be produced, for example
chlorinated organics, many of which are carcinogenic and/or teratogenic.
Quantitative microbial risk analysis (QMRA) permits health risks – here risks
from consuming wastewater-irrigated crops – to be calculated. The reference
level of tolerable risk is that adopted by the World Health Organization for a
Wastewater Re-use in Agriculture 247
Table 21.4 Physicochemical Quality of Three Waste Stabilization Pond
Effluents in Israel
Pond A
Pond B
Pond C
Chloride (C1-)
Bicarbonate (HCO3– )
Sulphate (SO32–)
Boron (B)
Phosphorus (P)
Sodium (Na)
Calcium (Ca)
Magnesium (Mg)
Note: Concentrations in mg/l, except SAR and pH
Source: Watson (1962)
person becoming ill with an excreta-related gastrointestinal infection (ie any of
those in Categories I and II, Chapter 2) as a result of drinking fully treated
drinking water. This tolerable risk is 10–3 per person per year, which means
that an individual has a 1 in 1000 chance of becoming ill, or one person in a
community of 1000 may become ill, from drinking fully treated drinking water
over a 12-month period (World Health Organization, 2003*).
Of course, carefully obtained epidemiological evidence is best to assess the
extent of actual health risks resulting from treated wastewater re-use. However,
if the calculated risks from consuming wastewater-irrigated crops are less than
10–3 per person per year, we can safely assume that the risks are acceptable. If
they are greater than 10–3 per person per year, then we might consider them too
high. However, Haas (1996) has argued that even 10–3 is too low. It is certainly
very low when compared with global diarrhoeal disease statistics: in developing
countries the actual incidence of diarrhoeal disease (rather than infection) is
~1.3 per person per year, and in ‘established market economy’ countries it is
~0.2 per person per year (Murray and Lopez, 1996b; see also Wheeler et al,
1999*); in other words, in developing countries everybody has diarrhoea at
least once a year, and even in industrialized countries individuals are at the high
annual diarrhoeal disease risk of ~1 in 5.
Only an outline introduction to QMRA is given here. Comprehensive
details are given by Haas et al (1999) and in Fewtrell and Bartram (2001*). A
general introduction to risk perception is given in Slovic (2000).
Dose-response models
There are two dose-response models currently used: the exponential model
and the β-Poisson model. Both are used to calculate first the risk of infection
which results from ingesting a single dose d of a microbial pathogen, and then
the annual risk resulting from multiple exposures to the dose d. Which model
248 Domestic Wastewater Treatment in Developing Countries
is used depends on the type of microbial pathogen: the exponential model is
used for the protozoa Cryptosporidium parvum and Giardia intestinalis (also
called G lamblia), and the β-Poisson model is used for the excreted viral and
bacterial pathogens.
Exponential model
The basic equation for the exponential dose-response model is:
PI(d) = 1 – exp(–rd)
where PI(d) is the probability of an individual, or the probable proportion of a
community, becoming infected after ingesting a singe dose of d excreted
protozoan pathogens; and r is a dimensionless ‘pathogen infectivity’ constant.
The value of r is 0.0042 for Cryptosporidium and 0.0199 for Giardia (Rose et
al, 1991; Haas et al, 1999). When rd is small, equation 21.2 becomes:
PI(d) = rd
Equation 21.2 can be rearranged to calculate the maximum dose
corresponding to any given level of risk:
D = – r–1ln[1 – PI(d)]
When PI(d) = 0.5 (ie 50 per cent of the community is infected), d becomes N50,
the median infectious dose. From equation 21.4:
N50 = – r–1ln(0.5) = 0.69/r
β-Poisson model
The basic equation for the β-Poisson dose-response model is:
PI(d) = 1 – [1 + (d/N50)(21/α – 1)]–α
where α is a dimensionless pathogen infectivity constant. Values of N50 and α
for selected excreted viral and bacterial pathogens are given in Table 21.5.
Equation 21.6 can be rewritten as:
d = {[1 – PI(d)]–1/α – 1}{N50/(21/α – 1)}
Equations 21.2–21.7 are for a single exposure to the pathogen dose d.
Wastewater Re-use in Agriculture 249
Table 21.5 Values of N50 and α for Selected Excreted Viral and Bacterial
Vibrio cholerae
Source: Haas et al (1999)
Multiple exposures
People regularly drink drinking water or eat wastewater-irrigated crops, so it
is necessary to be able to calculate the annual risk of infection resulting from
multiple, rather than single, exposures to the dose d.
The annual risk of infection per person from n exposures per year to a
pathogen dose d is denoted by PI(A)(d), which is given by:
PI(A)(d) = 1 – [1 – PI(d)]n
where PI(d) on the right-hand side of the equation is, as before, the risk of
infection from a single exposure to the pathogen dose d; and n is the number
of days in a year when a person is exposed to this single dose d. For drinking
water n is 365 as people drink water every day; for wastewater-irrigated crops
n could be 365 if the crops were eaten every day, but it could obviously be
less, for example, 365/2 if the crops were eaten every second day.
The terms on the right-hand side of equation 21.8 are explained as follows:
[1 – PI(d)]
[1 – PI(d)]n
1 – [1 – PI(d)]n
is the risk of not becoming infected from a single
exposure to the dose d;
is the risk of not becoming infected from n
exposures to the dose d; and
is therefore the risk of becoming infected from n
exposures to the dose d.
Infection and disease
Only a proportion of infected individuals will develop the disease:
PD(d) = γPI(d)
where PD(d) is the probability of disease in an individual exposed to a single
dose d; and γ is a constant with a value between 0 and 1.
250 Domestic Wastewater Treatment in Developing Countries
Worked example
Shuval et al (1997*) used QMRA to calculate the risks associated with the
consumption of salad crops irrigated with (a) untreated wastewater, (b)
wastewater treated to the WHO guideline level of ≤1000 E coli per 100 ml,
and (c) wastewater treated to the recommendation of the US Environmental
Protection Agency and the US Agency for International Development of zero
E coli per 100 ml (Environmental Protection Agency, 1992). They assumed
that a person eats 100 g of wastewater-irrigated lettuce every second day; there
is one virus per 105 E coli; and a 3-log pathogen die-off occurs between harvest
and consumption. Laboratory tests indicated that the mean volume of
wastewater remaining on 100 g of lettuce after irrigation was 10.8 ml.
Untreated wastewater
Untreated wastewater is taken to contain 107 E coli per 100 ml – that is, 105
per ml and therefore 10.8 x 105 E coli per 10.8 ml. This results in a count of
10.8 x 105 x 10–5 (ie 10.8) viruses per 100 g of lettuce at harvest, which is
reduced by the 3-log die-off to 1.08 x 10–2 virus per 100 g lettuce at
consumption. This is the single dose d to which an individual is exposed every
second day – that is on 365/2 days per year.
For rotavirus infection equations 21.6 and 21.8, with N50 = 6.2 and α =
0.253, yield:
PI(d) = 1 – [1 + (1.08 x 10–2/6.2) (21/0.253 – 1)]–0.253
= 6.3 x 10–3
PI(A)(d) = 1 – [1 – (6.3 x 10–3)]365/2
= 0.68
This is much higher than WHO’s tolerable risk of infection of 10–3 per person
per year (but lower than the actual incidence of diarrhoeal disease in
developing countries, which is ~1.3 per person per year).
Wastewater with 1000 E coli per 100 ml
The E coli and virus counts are now lower by a factor of 104 – that is, 100 g
of lettuce contains 1.08 x 10–6 virus. The equations yield:
PI(d) = 6.3 x 10–7
PI(A)(d) = 1.2 x 10–4
that is less than the tolerable risk of infection of 10–3 per person per year.
Wastewater Re-use in Agriculture 251
Wastewater with 1 E coli per 100ml
Using an E coli count of 1/100 ml (rather than 0/100 ml), the resulting virus
count per 100 g of lettuce is a thousand-fold less than in the second example –
that is, 1.08 x 10–9. The equations yield:
PI(d) = 6.3 x 10–10
PI(A)(d) = 1.2 x 10–7
that is excessively safe when compared with the tolerable risk of infection of
10–3 per person per year.
Shuval et al (1997*) calculated the additional costs of treating the
wastewater to one E coli per 100 ml, rather than to 1000 E coli per 100 ml,
and then the cost of each case of disease avoided: for rotavirus disease this was
US$3.5 millions and for hepatitis A US$35 millions! Clearly such costs can
never be justified, and the money would obviously be better spent on primary
health care facilities (even hospitals).
The above worked example, assumes a constant E coli–virus ratio and a
constant die-off between harvest and consumption, rather than a range of
values for each parameter and multi-trial Monte Carlo simulations.
Nevertheless it produces reasonably good order-of-magnitude estimates of risk.
QMRA studies in the US, which included Monte Carlo simulations, have
shown that the risks of viral infection were 10–3–10–5 per person per year for
consuming salad crops irrigated with a conventional (activated sludge)
effluent; effluent disinfection (in this case with chlorine) reduced the risks to
10–7–10–9 per person per year (Tanaka et al, 1998). Sleigh and Mara (2003b*)
detail a ‘freeware’ computer program for Monte Carlo QMRA.
As noted earlier in this chapter, crop irrigation with untreated wastewater is
common (see also Feenstra et al, 2000*; Ensink et al, 2002*; van der Hoek et
al, 2002*; and, more generally, International Water Management Institute,
2003*). It is, in fact, more common than irrigation with treated wastewater as
most wastewater in developing countries is not treated (Chapter 1) and many
farmers only have untreated wastewater with which to irrigate their crops. The
epidemiological evidence presented earlier in this chapter clearly shows that the
resulting health risks are very high, and this is confirmed by the QMRA study
of Shuval et al (1997*) referred to above. However, irrigation with untreated
wastewater will not, of course, cease – at least not in the short to medium term.
What, then, can be done to lessen these very high health risks?
This question was addressed in The Hyderabad Declaration on Wastewater
Use in Agriculture of 14 November 2002 (Resource Centre on Urban
Agriculture and Forestry, 2002*; see also Drechsel et al, 2002*). The
Declaration notes that ‘without proper management, wastewater use possesses
252 Domestic Wastewater Treatment in Developing Countries
serious risks to human health and the environment’, and recommends that,
until treatment becomes feasible, guidelines should be developed ‘for untreated
wastewater use that safeguard livelihoods, public health and the environment’,
together with the ‘application of appropriate irrigation, agricultural, postharvest and public health practices that limit risks to farming communities,
vendors and consumers’. Guidelines for the microbiological quality of treated
wastewaters used for crop irrigation cannot, of course, be relaxed, and this
means that the very high health risks resulting from irrigation with untreated
wastewater have to be managed, at least to some degree, by other means
(Figure 21.5; see also van der Hoek et al, 2002*). These include crop
restriction (ie only restricted irrigation), human exposure control (footwear to
avoid hookworm infection, and hand-washing to reduce bacterial, viral and
Ascaris infections), keeping very young children away from the rawwastewater-irrigated fields, and regular anti-helminthic chemotherapy (ie
regular ‘deworming’) for fieldworkers and any exposed children.
Wastewater treatment should be gradually phased in, for example, primary
treatment (in, for example, short-retention-time anaerobic ponds) should be
implemented first as it is so very effective in reducing nematode egg counts
(Chapter 11) and thus the health risks posed by these nematodes, and this is
clearly much better than no treatment at all. Of course, municipal authorities
may be reluctant or unable to invest in even primary treatment, but anaerobic
ponds can fairly easily be excavated by farmers on a small part of their land
(or they can group themselves into a cooperative for this purpose). The heath
risks from irrigation with untreated wastewater can be managed, at least to
some extent, but the farmers need to understand the risks, how their practices
(eg not wearing anything on their feet, poor personal hygiene) aggravate these
risks, and how relatively simple things can be done to reduce them. Diarrhoeal
disease (due to all transmission routes, not just raw-wastewater irrigation) is
likely to be high in poor farming communities, so interventions to improve
water supplies, sanitation and hygiene will be required, in addition to specific
improvements in irrigation practices, before any real health improvements
become apparent.
Wastewater Re-use in Aquaculture
Aquaculture means ‘water farming’, just as agriculture means ‘field farming’,
and so it encompasses fish culture and growing aquatic vegetables. The fish
most frequently grown in aquaculture ponds are carp and tilapia. Local species
should be grown – in India, for example, Indian major carp are mainly grown,
such as catla (Catla catla), mrigal (Cirrhina mrigala) and rohu (Labeo rohita).
Kolkata (formerly Calcutta) provides the world’s largest example of
wastewater-fed fisheries (Jana, 1998*; Nandeesha, 2002*): around 3500 ha of
fishponds are fertilized with 550,000 m3/day of untreated wastewater (Figure
22.1), and fish production (Indian major carp, with some tilapia and silver
carp) is ~20 tonnes/day, equivalent to ~18 per cent of the city’s demand for
fish. The average fish yield is 4 t/ha year, although the better managed ponds
produce 7–8 t/ha year. Some 17,000 local people work on the fishponds
(Edwards, 2001*), and the fish sell for around US$1/kg – it is a highly
profitable business.
In China and Indonesia where wastewater-fed (or excreta-fertilized)
aquaculture has been practised for many hundreds of years, several carp species
are often grown in the same pond (a practice known as ‘polyculture’), with
each species occupying a different ecological niche. The fish grown include
bighead carp (Aristichthys nobilis), silver carp (Hypophthalamichthys molitrix),
grass carp (Ctenopharyngodon idellus), and common carp (Cyprinus carpio).
Figure 22.1 Some of the Kolkata East Wastewater-fed Fishponds
254 Domestic Wastewater Treatment in Developing Countries
Figure 22.2 Harvesting Indian Major Carp from the Kolkata East
Wastewater-fed Fishponds
Wastewater Re-use in Aquaculture 255
Tilapia are widely cultured throughout the developing world – they are even
called ‘aquatic chickens’ as they are so easy to produce and they provide highquality animal protein at an affordable price in many countries (Coward and
Little, 2001*). The most commonly cultured species are Nile tilapia
(Oreochromis niloticus) and Mozambican tilapia (O mossambicus). An
improved strain of O niloticus has been developed through selective breeding
by the World Fish Center – the so-called GIFT (genetically improved farmed
tilapia) strain (Pullin et al, 1991; ICLARM, 2000). This grows ~60 per cent
faster than other strains and can be harvested three times a year rather than
twice. As yet there has been almost no work on growing the GIFT strain in
wastewater-fed fishponds; clearly, given the popularity of tilapia as a highprotein food, this is an area of research urgently needed. Some initial work has
been done in Ho Chi Minh City, Vietnam: polluted urban canal water, which is
essentially septic tank effluents and untreated industrial wastewaters, is batchfed into a shallow pond and left for seven days, when GIFT brood fish are
introduced; the resulting young fish are raised to fingerling size after five weeks;
they are then harvested and sold to local fish farmers for ~US$4/kg. A 1000 m2
pond produces ~200 kg of fingerlings every 6–8 weeks. The fish farmers then
raise the fingerlings intensively in freshwater ponds, using industrialized fish
food, to marketable size (~250–350 g) within four months (Minh, 2002).
In Asia particularly, a wide range of aquatic vegetables is grown (not just
watercress as in most industrialized countries), for example, water spinach
(Ipomoea aquatica), water chestnut (Eleocharis dulcis and E tuberosa), water
bamboo (Zigania spp), water calthrop (Trapa spp) and lotus (Nelumbo
Extensive reviews of wastewater-fed aquaculture are given by Edwards
and Pullin (1990) and Edwards (1992*, 1999*). Yan et al (1998*) note that
wastewater-fed aquaculture complies with the fundamental Chinese
philosophical principles of ‘holism, harmony, self-resiliency, regeneration and
circulation’, and that its objectives are the conversion of wastes into usable
resources (ie food), environmental protection (pollution being avoided by
waste utilization), and sustainable development.
Hoan (1996) gives interesting data for Hanoi: in the Thanh Tri district of
the city, the main wastewater-re-use area, a polyculture of tilapia, silver carp
and two Indian major carp (mrigal and rohu) grown in wastewater-fed ponds
yields ~5.6 t/ha year, compared with ~4.1 t/ha year from non-wastewater-fed
ponds – a ~36 per cent increase due to the use of wastewater. There is a much
greater profit (ie fish sales less all operating costs) from wastewater-fed ponds:
US$1000/ha/year, compared with US$300 from non-wastewater-fed ponds – a
230 per cent increase, due in part to the higher fish yield but also because
expensive fish food does not have to be purchased.
There has been a large research effort in Lima, Peru on growing tilapia in
wastewater-fed ponds, including in the final maturation ponds at San Juan
WSP site. Yields of fish grown in these maturation ponds (without
supplementary feeding) are ~4.5 tonnes/ha per harvest, with two harvests per
year (Cavallini, 1996*; Nava, 2001).
256 Domestic Wastewater Treatment in Developing Countries
As noted above, treated wastewater can be profitably used to fertilize
fishponds and aquatic vegetable ponds. As with crop irrigation with treated
wastewater (Chapter 21), the treated wastewater used for aquaculture must be
of a minimum microbiological quality to ensure that the products of
aquaculture (fish, aquatic vegetables) are safe and that there is no excess risk
of infection to the aquaculture pond workers.
Microbiological quality guidelines
Human nematode infections (which are important in the case of wastewater
re-use in agriculture) are not important in aquacultural re-use. Rather, it is the
human trematode infections that are of major concern; these are the Category
V water-based helminthic infections (Chapter 2), and the most important of
these are:
the human schistosomes or blood flukes, principally Schistosoma mansoni,
S haematobium and S japonicum,
the Oriental liver fluke, Clonorichis sinensis, and
the giant intestinal fluke, Fasciolopis buski.
All these trematodes have an aquatic snail as their primary intermediate host.
The schistosomes do not have a secondary host, but Clonorchis has fish
(usually members of the carp family) and Fasciolpsis has aquatic vegetables as
secondary hosts. Thus, in endemic areas, the schistosomes present a risk to the
aquaculture pond workers, and the other two to the consumers of the
aquaculture products. Cancer of the bile duct (‘cholangiocarcinoma’) is a not
uncommon outcome of chronic clonorchiasis (Chapter 2; see also Wiwanitkit,
2003*) and this disease usually has a high prevalence whenever the local
people eat uncooked fish grown in wastewater-fed fishponds.
Because these trematodes undergo massive asexual multiplication in
aquatic snails (Chapter 2), the World Health Organization (1989*)
microbiological quality guideline for aquaculture is:
zero viable human trematode eggs per litre of treated wastewater.
The consumers also need protection against Category II bacterial diseases
(Chapter 2), and the WHO guideline for this is:
≤1000 E coli per 100 ml of aquaculture pond water.
If the E coli count is above this guideline level, then there is a risk that bacterial
pathogens will be present, not only on the skin of the fish, but also in their
flesh and internal organs (Buras et al, 1987*; Pal, 1991). However, Cavallini
(1996*), who investigated Nile tilapia grown in fishponds fed with waste
Wastewater Re-use in Aquaculture 257
stabilization pond effluents at the San Juan WSP site in Lima, Peru, found that
pathogen-free fish were obtained when the fishpond waters contained up to
104 E coli per 100 ml – so there is a margin of safety in the WHO guideline
(see also World Health Organization, 1995, 1999*). Hepher et al (1986) found
that E coli die-off was higher when fish (tilapia) were present than when they
were absent. [Note added in proof: The 2004 revised WHO bacterial guideline
for E coli numbers in wastewater-fed fish ponds will be ≤104/100 ml, rather
than ≤1000/100 ml as in the 1989 guidelines – a relaxation of an order of
magnitude (see WHO, 2004*). In practice (and as shown in the Design
Example on p260) this will make little or no difference to waste-water-fed
fishponds designed on the basis of a total N loading of 4 kg/ha day.]
Buras et al (1987*) proposed that the sanitary quality of fish grown in
wastewater-fed ponds should be assessed by the bacterial count per gram of
fish muscle (ie flesh). For numbers of both total aerobic bacteria (ie the
‘standard plate count’ on nutrient agar at 35–37°C) and E coli the quality is
interpreted as follows:
0–10 per g:
11–30 per g:
31–50 per g:
>50 per g:
Very good
The design procedure given here for wastewater-fed fishponds is based on the
concept of minimal wastewater treatment for maximal production of
microbiologically safe fish (Mara et al, 1993; Mara, 1997*). The design steps
are as follows:
Design an anaerobic pond and a secondary facultative pond, as detailed in
Chapters 10 and 11.
Use equation 12.12 to determine the total nitrogen concentration in the
facultative pond effluent (Ce, mg N/1).
Design the wastewater-fed fishpond, which receives the facultative pond
effluent, on the basis of a surface loading of total nitrogen (λsTN) of 4 kg/ha
day. (Too little nitrogen results in a low algal biomass in the fishpond and
consequently small fish yields. Too much nitrogen gives rise to high
concentrations of algae, with the resultant high risk of severe dissolved
oxygen depletion at night and consequent fish kills. A loading of ~4 kg
total N/ha day is optimal.)
The fishpond area (Afp, m2) is given by the following version of equation
Afp = 10CeQ/λsTN
258 Domestic Wastewater Treatment in Developing Countries
Use equation 11.7 to calculate the retention time in the fishpond (θfp,
days), with a fishpond depth of 1 m.
Use the following version of equation 12.3 to calculate the number of E
coli per 100 ml of fishpond water (Nfp):
Nfp = Ni/(1 + kTθa) (1 + kTθf) (1 + kTθfp)
Nfp should be ≤1000 per 100 ml. If it is not, increase θfp until it is (or
consider having a maturation pond ahead of the fishpond).
Use equation 12.14 or equation 12.15 to determine the concentration of
NH3-N first in the facultative pond effluent (assume that the conversion of
organic nitrogen in the anaerobic pond to ammonia produces an ammonia
concentration in the effluent of the anaerobic pond – that is, in the influent
to the facultative pond – equal to 75 per cent of the total nitrogen
concentration in the raw wastewater), and then in the fishpond.
The ammonia concentration is the total concentration of NH3 and
NH 4 (ie ‘free and saline’ ammonia). In order to protect the fish from free
ammonia (NH3) toxicity, the concentration of NH3 should be <0.5 mg
N/1. The percentage (p) of free ammonia in aqueous solutions depends on
temperature (T, K) and pH, as follows (Emerson et al, 1975; see also
Erickson, 1985):
p = 100[10(pKa–ph) + 1]
where pKa is given by:
pKa = 0.09018 + (2729.92/T)
where T is the absolute temperature in degrees Kelvin (K =°C + 273).
Equations 22.3 and 22.4 (or Table 22.1 which is derived from them)
should be used to determine the free ammonia concentration in the fishpond,
assuming a pH of 7.0 (the pH range in wastewater-fed fishponds is usually
Fish yields
Good fishpond management can be achieved by having small ponds, generally
≤1 ha, that can be stocked with fingerlings at the rate of 3/m2, fertilized with
facultative pond effluent and then harvested 3 months after stocking. During
this time the fingerlings will have grown from ~20 g to ~200 g. Partially
draining the pond will ensure that almost all the fish can be harvested. This
Wastewater Re-use in Aquaculture 259
Table 22.1 Percentage of Free Ammonia (NH3) in Aqueous Ammonia
(NH3+NH4) Solutions at 10–25 °C and pH 7.0–8.5
Temperature (ºC)
Percentage of free ammonia in aqueous ammonia
solutions at pH
Source: Emerson et al (1975)
cycle can be done two or three times per year. Allowing for a 30 per cent fish
loss due to mortality, poaching and consumption by fish-eating birds, the
annual yield is:
(3 x 200 g fish per m2) (10–6 tonnes/g) (104 m2/ha) x (2–3 harvests per year) x
(0.7, to allow for the 30 per cent loss)
= 8–12 tonnes of fish per hectare per year.
If there is a high local demand for fish, then it is often sensible to use treated
wastewater in both agriculture and aquaculture (Food and Agriculture
Organization, 2001). In practice this means first use the treated wastewater to
grow fish and/or aquatic vegetables, and then use the effluent from the
aquaculture ponds for crop irrigation. In Kolkata, for example, the fishpond
effluent is used to cultivate rice; alternatively, fish can be grown in rice paddies
(see Berg, 2002*). In the Zhujiang delta area in southern China crops such as
sugar cane and mulberry bushes are grown on the fishpond embankments, and
rice paddies are fertilized with fishpond effluents, in this centuries-old model
of intensive integrated agriculture and aquaculture (Ruddle and Zhong, 1988).
260 Domestic Wastewater Treatment in Developing Countries
Design a wastewater-fed fishpond system to receive the effluent from the
secondary facultative pond designed in Chapter 11. Take the total nitrogen
and ammonia concentrations in the raw wastewater as 50 mg N/l and 30 mg
N/l, respectively.
The anaerobic pond has an area of 3340 m2 and a retention time of 1 day
(Chapter 10), and the secondary facultative pond has an area of 26,500 m2
and a retention time of 4 days (Chapter 12).
Fishpond area
Assuming no total nitrogen removal in the anaerobic pond and a pH of 7.5 in
the facultative pond, the total nitrogen concentration in the effluent from the
facultative pond is determined from equation 12.12:
Ce = Ci exp{– [ 0.0064(1.039)T–20][θf + 60.6(pH – 6.6)]}
= 50 exp{– [0.0064(1.039)25–20][4 + 60.6(7.5 – 6.6)]}
= 32 mg total N/l.
Using equation 22.1 the fishpond area is:
Afp = 10CeQ/λsTN
= 10 x 32 x 10,000/4
= 800,000 m2 (80 ha)
Using equation 22.2, determine the number of E coli per 100 ml of fishpond
Nfp = Ni/(1 + kTθa)(1 + kTθf)(1 + kTθfp)
The retention time in the fishpond is given by equation 11.7:
θfp = 2AfpDfp/(2Q – 0.001eAfp)
With Dfp = 1 m and e = 5 mm/day:
θfp = (2 x 800,000 x 1)/[(2 x 10,000) – (0.001 x 5 x 800,000)]
= 100 days.
Wastewater Re-use in Aquaculture 261
E coli numbers
kB(T) = 2.6 (1.19)T–20
= 6.2 d–1 for T = 25°C
Nfp = 5 x 107/[1 + (6.2 x 1)][1 + (6.2 x 4)][1 + (6.2 x 100)]
= 430 per 100 ml (satisfactory as <1000)
Free ammonia
As T = 25°C, use equation 12.15 to determine the concentration of total
ammonia in the facultative pond effluent:
C = Ci/{1 + [5.035 x 10–3(A/Q)][exp(1.540 x (pH – 6.6))]}
With Ci = 0.75 x 50 = 37.5 mg/l and pH = 7.0:
C = 37.5/{1 + [5.035 x 10–3(26 500/10,000)][exp(1.540 x (7.0 – 6.6))]}
= 36.6 mg/l
Note: this value is higher than the total N concentration determined above for
the facultative pond effluent. This is an inherent problem with two empirical
equations derived from different data sets. However, it is a conservative error
since the free ammonia concentration is slightly overestimated.
The percentage of free ammonia is given by equations 22.3 and 22.4:
pKa = 0.09018 + (2729.92/T)
= 0.09018 + [2729.92/(25 + 273)]
= 9.25
p = 100[10(pKa–pH) + 1]–1
= 0.566 per cent
Thus the concentration of free ammonia in the fishpond is 0.00566 x 36.6 =
0.21 mg N/l (satisfactory as <0.5 mg N/l).
Pond areas
• Anaerobic pond
• Facultative pond
• Fishpond
• Total
3340 m2
26,500 m2
800,000 m2
829,840 m2
Thus only 3.6 per cent of the total area is used for wastewater treatment.
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Journal URLs
The following journals are available at Science Direct (http://www.sciencedirect.com):
Chemical Engineering Science
Critical Reviews in Environmental Science and Technology
Critical Reviews in Microbiology
Ecological Economics
Ecological Engineering
Environment International
References 287
Microbes and Infection
Parasitology Today
Urban Water
Water Policy
Water Research
The URLs of other journals are:
Annual Review of Microbiology: http://micro.annualreviews.org
Applied and Environmental Microbiology: http://aem.asm.org
British Medical Journal: http://www.bmj.com
Bulletin of the World Health Organization: http://www.who.int/bulletin/index.htm
Clinical Microbiology Reviews: http://cmr.asm.org
Communicable Disease and Public Health: http://www.phls.org.uk/publications/
Emerging Infectious Diseases: www.cdc.gov/ncidod/EID/index.htm
Environmental Health Perspectives: http://ehp.niehs.nih.gov
Environmental Management: http://link.springer.de/link/service/journals/00267/
International Journal of Environmental Health Research: http://www.tandf.co.uk/
Journal of the American Water Resources Association: http://www.awra.org/jawra
Journal of Applied Microbiology: http://www.blackwellpublishing.co./journals/jam/
Journal of Environmental Engineering, American Society of Civil Engineers:
Journal of Water and Health: http://www.iwaponline.com/jwh/toc.htm
The Lancet: http://www.thelancet.com
Policy Review: http://www.policyreview.org
Proceedings of the National Academy of Sciences: http://www.pnas.org
Tropical Medicine and International Health: http://blackwellpublishing.com/
Urban Agriculture Magazine: http://www.ruaf.org/newslgeneng.html
Water, Air and Soil Pollution: http://www.kluweronline.com/issn/0049-6979/current
Waterlines: http://isacco.ingentaselect.com/vl=79944246/cl=22/nw=1/catchword/
itpub/ 02628104/contp1-1.htm
Water Science and Technology: http://www.iwaponline.com/wst/toc.htm
Water Supply: http://www.iwaponline.com/ws.toc.htm
IWA waste stabilization pond conference proceedings
The International Water Association has organized six conferences on waste
stabilization ponds, the proceedings of which are published in Water Science and
Technology (WST). All are available on-line, except those of the first international
First international conference (Lisbon, 1987): WST, 1987, vol 19, no 12
Second international conference (Berkeley, 1993): WST, 1995, vol 31, no 12
Third international conference (João Pessoa, 1995): WST, 1996, vol 33, no 7
Fourth international conference (Marrakech, 1999): WST, 2000, vol 42, no 10–11
First Latin American regional conference (Cali, 2000): WST, 2002, vol 45, no 1
Fifth international conference (Auckland, 2002): WST, 2003, vol 48, no 2
Sixth international conference (Avignon, 2004): WST, 2005 (in press)
Details of further conferences in this series will be available at http://www.
iwahq.org.uk (click on ‘Events’).
288 Domestic Wastewater Treatment in Developing Countries
IWA wastewater re-use conferences
The IWA has organized six conferences on wastewater reclamation, recycling and reuse,
the proceedings of which are published in Water Science and Technology (WST) or
Water Supply (WS). They are all available on-line, except those of the first conference:
First international conference (Costa Brava, 1991): WST, 1991, vol 24, no 9
Second international conference (Iraklio, 1995): WST, 1996, vol 33, no 10–11
First Mediterranean regional conference (Milan, 1998): WST, 1999, vol 40, no 4–5
Third international conference (Paris, 2000): WST, 2001, vol 43, no 10
Second Mediterranean regional conference (Iraklio, 2002): WS, 2003, vol 3, no 4
Fourth international conference (Mexico City, 2003): WST, 2004 (in press)
Details of further conferences in this series will be available at http://www.iwahq.org.uk
(click on ‘Events’).
Sanitation Connection
Sanitation Connection is a directory of publications on low-cost sanitation in
developing countries, including appropriate wastewater treatment and wastewater
reuse in agriculture and aquaculture, which are available on the Internet. Its URL is
http://www.sanicon.net/home.php3; click on the topic required and then on
‘publications’; each publication has its own page with a link to the on-line document.
Topic coordinators do their best to keep the list of publications up-to-date, so it is a
very useful resource.
acceptable risk, 247
activated sludge, 71
actual health risks, 232
advanced pond systems, 102–104
aerated lagoons, 71, 101, 213–225
field performance, 217–219
floating, 215
agricultural re-use, 42–43, 230–252
agricultural–aquacultural re-use,
integrated, 259–260
Al Samra ponds, 95–96, 171
algae, 37, 114–116,
algal biomass, 125–130
algal productivity, 125–126
algal–bacterial mutualism, 86
ammonia, 52, 149–150, 198
toxicity, 130, 258
amoebiasis, 9
anabolism, 28–29
anaerobic digestion, 31–33
anaerobic ponds, 85, 105–113, 147,
by-pass pipework 169–170
cover, 169–170
depth, 108
design, 108–109
desludging, 176–179
E coli removal, 147
high-rate, 110–111
in series, 110–112
sludge accumulation, 109–110
Ancylostoma duodenale, 14
aquacultural re-use, 42–43, 253–262
aquatic vegetables, 256
Archaea, 21–35
Arrhenius constant, 59–60
Aruba protocol, 53–54
ascariasis, 9
Ascaridia galli, 236–237
Ascaris lumbricoides, 10, 11, 14, 140,
autolysis, 28–29
Avogadro’s number, 127
Bacteria, 21–35
bacterial growth
curve, 26–27
growth kinetics, 26–29
bacteriochlorophylls, 132
bacteriophages, 24
Bancroftian filariasis, 9
beta-Poisson dose-response model,
biochemical oxygen demand – see BOD
biofilters, 71, 207–212
biotic index, 38–40
BOD (biochemical oxygen demand),
3–4, 49–51, 85, 87–88
5-day, 57–58
curves, 57
filtered, 51, 121
removal kinetics, 56–68
surface loading, 114, 118–120, 128
ultimate, 57–58, 66–67
volumetric loading, 108–109
boron, 244
bubbler irrigation, 241
Campylobacter, 13,
campylobacteriosis, 9
cancer, 16
carp, 253
Indian major, 253
Cartagena convention, 53
catabolism, 28–29
chemical oxygen demand – see COD
Chlamydomonas, 105, 114, 129, 130
290 Domestic Wastewater Treatment in Developing Countries
Chlorella, 114, 129, 130
Chlorobiaceae, 130–132
chlorophyll a, 125, 129
cholera, 9
clonorchiasis, 9
Clonorchis sinensis, 16, 43, 256
COD (chemical oxygen demand), 3
coliform bacteria, 33–34
column sampler, 186
complete mixing, 58–60, 86–88
composite exponential, 64–65
conductivity, 243
constant velocity grit channels, 81–83
constructed wetlands, 71, 100–101,
continuous culture, 27
Cromatiaceae, 130–132
crop restriction, 241
crop yields, 230
cryptosporidiosis, 9
Cryptosporidium parvum, 12, 140–141
Culex quinquefasciatus, 9
Cyclospora cayentanensis, 10
Dandora ponds, 95
data storage and analysis, 186–187
decentralized treatment, 70–71
denitrification, 31
dispersed flow, 60–62
dispersion number, 61–62
dissolved oxygen
diurnal variation in ponds, 115
sag curve, 44–47
domains of life, 21–22
dose–response models, 247–249
drip irrigation, 241
E coli, 9, 13, 34–35, 42–43, 52,
237–239, 256
enteropathogenic, 13
removal equations, 125, 141–148
effluent discharge
coastal waters, 52–54
inland waters, 43–47
effluent quality, 41–55, 94, 121–122,
effluent standards, 48–54
effluent take-off levels, 168
electrical conductivity, 243
protection, 160–162
slope, 159
emerging infectious diseases, 16
Entamoeba histolytica, 12, 237
enterobiasis, 9
Escherichia coli – see E coli
Euglena, 114, 129, 130
Eukarya, 21–22
excreta-related diseases, 8–19
cancers, 17–18
environmental classification, 8–17
global burden, 18–19
excreted load, 10
exponential dose–response model, 248
facultative ponds, 85–86, 114–135
algae, 116
BOD removal, 120–121
depth, 120
design, 118–125
E coli removal
function, 115
mixing, 115–117
stratification, 115–117
faecal coliforms – see E coli
faecal indicator bacteria, 33–35
faeces, 1
faeco-oral diseases, 9, 12–14
fasciolopsiasis, 9
Fasciolopsis buski, 16, 43, 256
filariasis, 9, 16
first-order kinetics, 56–64,
fish yields, 258–259
fishponds, 257–259
flow measurement, 84
flow regimes, 60–62
fly control, 176, 211
freeboard, 165
freshwater biology, 38–40
future projections, 77
geohelminthiases, 9, 14, 19
geotechnical aspects, 158–162
Giardia intestinalis, 12, 140–141, 237
Giardia lamblia – see Giardia intestinalis
giardiasis, 9
Gram stain, 24
disposal, 84
removal, 81–84
separators, 84
Index 291
gross areal oxygen production, 126
growth curve, 26–27
growth kinetics, 26–29
health risks
actual, 232
potential, 232
Helicobacter pylori, 16
helminth eggs, 42–43, 198, 233–237
removal equation, 124
helminths, 37
Henry’s law, 106
hepatitis, 9
high-rate algal ponds, 102–103
hookworms, 9, 140
human exposure control, 241–242
humus removal, 212
hybrid pond-reservoir system, 190
hydraulic flow regimes, 60–62
hydraulic loading rate, 152
hydrogen sulphide, 106–107
hymenolepiasis, 9
incidence, 12
industrial wastewaters, 75, 112, 245
infection, 12
infectivity, 10
infiltration, 75
inlet structures, 166–168
insect-vector diseases, 9, 16
intestinal nematode eggs – see helminth
irradiance, 127
localized (bubbler), 241
localized (drip), 241
restricted, 232
unrestricted, 232
untreated wastewater, 233–235,
land prices, 72
latency, 10
Leptospira interrogans, 16
leptospirosis, 9
light intensity, 126–128, 140
light-and-dark-bottle test, 125–126
lime-assisted sedimentation, 71–72
lining of ponds, 162
localized irrigation, 241
location of ponds, 158
logarithmic growth, 26
macrophyte ponds, 101–102
Mangere ponds, 98–100
Marais’ theorem, 88
maturation ponds, 85–86, 136–157, 220
bacterial removal, 138–140
BOD removal, 1481–49
depth, 136
design, 141–151
E coli removal, 141–148
function, 136
helminth egg removal, 140–141
pathogen removal mechanisms,
viral removal, 137–138
von Sperling’s equations
McGarry and Pescod equation, 118–119
membrane bioreactors, 72
microbiology, 20–35
micro-invertebrates, 38–40
micro-organisms, 20–22
mid-depth area, 164–165
minimal water quality index
Monod equation, 28
multiple exposures, 249
multiplication, 10
mutualism, 86
Necator americanus, 14
nematode eggs – see helminth eggs
nitrification, 30–31, 151
nomenclature, 21
norovirus, 9, 22
nutrient removal, 149–151, 216–217
odour, 106–107
Oreochromis, 255
outlet structures, 166–169
oxidation ditches, 71, 225–229
oxypause, 114
Pano and Middlebrooks equations,
partial treatment, 241
peak wastewater flow, 76
permeability, coefficient of, 162
persistence, 11
pH, 25, 107, 114, 139, 245
292 Domestic Wastewater Treatment in Developing Countries
phase separators, 204–205
phosphorus removal in ponds, 151
photo-inhibition, 128
photon flux density, 127
photosynthesis, 37, 125–126
photosynthetic bacteria, 130–132
photosynthetic bacteria, 31
photosynthetically active radiation, 127
physicochemical river water quality, 48
Planck’s constant, 127
plug flow, 60, 86–88
pond effluent polishing, 151–152
pond geometry, 163–164
potential health risks, 232
preliminary treatment, 78–84
prevalence, 12
Proctor test, 159
protozoa, 35–36, 237–238
protozoan cysts, 140–141
purple ponds, 130–132
Pyrobotrys, 114
QMRA, 11, 246–251
quantitative microbial risk analysis – see
Reed’s equation, 149
restricted irrigation, 232
retarded exponential, 65–66
re-use – see wastewater re-use
actual, 232
potential, 232
risk analysis – see QMRA
rock filters, 151–152
rodent-vector diseases, 9, 17
rotavirus, 9, 22
Salmonella, 13,
salmonellosis, 9
Scenedesmus, 129, 130
Schistosoma, 16, 43, 256
schistosomiasis, 9, 10
screening, 78–80
scum guard, 167–168
sedimentation pond, 220–222
Shigella, 13
shigellosis, 9
sludge accumulation, 109–110
sludge drying beds, 110, 204
sludge layer, 118
sodium adsorption ratio, 243
specific growth rate, 26
Streeter–Phelps equation, 45–47
strongyloidiasis, 9
sullage, 1
sulphates, 106–107
sulphides, 106–107
toxicity, 128–129
susceptibility, 12
suspended solids, 5, 94, 196
sustainability, 69–70
Taenia saginata, 14
Taenia solium, 14
taeniasis, 9, 14
temperature, 25, 59–60, 109, 139
design, 109
theoretical oxygen demand, 2–3
tilapia, 255,
tolerable risk, 247
total nitrogen, 244,
removal in ponds, 149
treatment objectives, 41–42
tree of life, 22–23
treebelt, 171
trematode eggs, 43, 256
trichuriasis, 9
Trichuris trichiura, 14, 140
typhoid, 9
UASBs, 71, 101
ultimate BOD, 57–58
uncertainty, 121–124
unrestricted irrigation, 232
upflow anaerobic sludge blanket reactors
– see UASBs
urine, 1
Vibrio cholerae, 13, 25, 85, 107
viruses, 22–24, 137–138
waste stabilization ponds, 71, 85–187
advantages, 89–92
disadvantages, 93–94
evaluation, 183–186
functions, 85–86
high altitude, 100
large systems, 95–100
layout, 86–87
Index 293
monitoring, 182–183
operation and maintenance, 175–181
physical design, 158–174
rehabilitation, 180
staffing levels, 179
start-up, 175
types, 85–86
upgrading, 173–174
usage, 94–100
application, 241
collection, 5
flows, 74–76
loads, 77
wastewater re-use, 42–43, 230–262
guidelines, 42–43
Islamic countries, 231
wastewater storage and treatment
reservoirs, 71, 188–193
wastewater strength, 4–5
wastewater treatment
investment, 5–6
options, 69–73
water-based helminthiasies, 14
Wehner–Wilhelm equation, 61
Werribee ponds, 96–98
white towel test, 176–177
WHO guidelines – see wastewater re-use
wind aerators, 132
wind mixing, 115–117
yersiniosis, 9
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