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Possibilities for sustainable nutrient recycling by faecal separation combined with urine diversion /



Household wastewater can be divided into three fractions by origin; urine, faeces and greywater. The largest nutrient and smallest heavy metal contents are found in the urine, which is easily collected separately using a urine-diverting toilet. The second most nutrient-containing fraction is the faecal matter. This fraction (faeces and toilet paper) has the smallest mass of the three, approximately 60 kg of wet weight per person and year. The nutrients in the urine and faeces have to be recycled to agriculture for society to be sustainable. The faecal matter can either be collected dry or, after a short waterborne transport, be separated from the flushwater in a separator that uses a combination of whirlpool effect, gravity and surface tension. Using this type of separation, between 58% and 85% of the faecal nutrients were separated in the measurements performed here. By recycling the urine and the faecal nutrients, much energy can be saved as the load on the wastewater treatment plant decreases and as mineral fertilisers are replaced in agriculture. To avoid transmission of diseases, the faecal matter has to be sanitised before recycling. If the faecal matter is collected dry, it is possible to perform the sanitation by thermal composting, preferably together with household biodegradable waste. A calculation method for determination of the safety margins for sanitation was developed. In a pilot-scale study, the safety margin for thermal composting of faeces and food waste, with old compost as an amendment, was approximately 37 times total inactivation of Enteroviruses, the most thermotolerant organism evaluated. Another sanitation method investigated was chemical disinfection using urea or peracetic acid. At a dosage between 0.5% and 1.0%, the highly reactive peracetic acid inactivated all investigated organisms within 12 hours of treatment. The high dry matter content (10% DM) meant that high dosages were needed. Lower dry matter content would decrease the dosage required for proper sanitation. A very promising treatment was the addition of urea. Addition of 30 g ureanitrogen per kg of wet weight faecal matter resulted in total inactivation of the monitored organisms, E. coli, Salmonella spp, Enterococcus spp, Salmonella typhimurium 28B phage and Ascaris suum eggs, within 50 days of treatment at 20°C. The spore-forming bacteria Clostridium spp in its dormant state was resistant to this treatment. As the urea has to be degraded to ammonia before it functions as a disinfectant, there is some delay in this treatment. Therefore, urea addition followed by 2 months storage is the preferred treatment for disinfection of separated faecal matter. As additional effects, urea increases the fertiliser value of the treated material and there is no risk of microbial regrowth. Changing to urine-diversion combined with faecal separation and disinfection by urea seems to be an interesting way to decrease the resource usage and possibly improve the hygienic standard of wastewater systems.
Household wastewater can be divided into three fractions by origin; urine, faeces
and greywater. The largest nutrient and smallest heavy metal contents are found in
the urine, which is easily collected separately using a urine-diverting toilet. The
second most nutrient-containing fraction is the faecal matter. This fraction (faeces
and toilet paper) has the smallest mass of the three, approximately 60 kg of wet
weight per person and year. The nutrients in the urine and faeces have to be
recycled to agriculture for society to be sustainable.
The faecal matter can either be collected dry or, after a short waterborne transport,
be separated from the flushwater in a separator that uses a combination of
whirlpool effect, gravity and surface tension. Using this type of separation,
between 58% and 85% of the faecal nutrients were separated in the measurements
performed here. By recycling the urine and the faecal nutrients, much energy can
be saved as the load on the wastewater treatment plant decreases and as mineral
fertilisers are replaced in agriculture.
To avoid transmission of diseases, the faecal matter has to be sanitised before
recycling. If the faecal matter is collected dry, it is possible to perform the
sanitation by thermal composting, preferably together with household
biodegradable waste. A calculation method for determination of the safety margins
for sanitation was developed. In a pilot-scale study, the safety margin for thermal
composting of faeces and food waste, with old compost as an amendment, was
approximately 37 times total inactivation of Enteroviruses, the most thermo-
tolerant organism evaluated.
Another sanitation method investigated was chemical disinfection using urea or
peracetic acid. At a dosage between 0.5% and 1.0%, the highly reactive peracetic
acid inactivated all investigated organisms within 12 hours of treatment. The high
dry matter content (10% DM) meant that high dosages were needed. Lower dry
matter content would decrease the dosage required for proper sanitation.
A very promising treatment was the addition of urea. Addition of 30 g urea-
nitrogen per kg of wet weight faecal matter resulted in total inactivation of the
monitored organisms, E. coli, Salmonella spp, Enterococcus spp, Salmonella
typhimurium 28B phage and Ascaris suum eggs, within 50 days of treatment at
20°C. The spore-forming bacteria Clostridium spp in its dormant state was
resistant to this treatment. As the urea has to be degraded to ammonia before it
functions as a disinfectant, there is some delay in this treatment. Therefore, urea
addition followed by 2 months storage is the preferred treatment for disinfection of
separated faecal matter. As additional effects, urea increases the fertiliser value of
the treated material and there is no risk of microbial regrowth.
Changing to urine-diversion combined with faecal separation and disinfection by
urea seems to be an interesting way to decrease the resource usage and possibly
improve the hygienic standard of wastewater systems.
Key words: Faecal separation, urine-diversion, nutrient recycling, wastewater
composition, wastewater reuse, disinfection safety margins, sanitation,
disinfection, thermal composting, chemical disinfection, urea, peracetic acid
Hushållsavloppsvattnet kan utifrån dess källor delas upp i tre fraktioner: urin,
fekalier och BDT vatten (bad-, disk-, tvättvatten). Urinen innehåller den största
delen växtnäring samtidigt som den innehåller de lägsta halterna
tungmetallföroreningar. Urinen samlas lätt upp separat i en urinsorterande toalett.
Fekalierna är den fraktion som innehåller näst mest växtnäring. Fekaliefraktionen
(fekalier och toalettpapper) är mycket liten, ungefär 60 kg per person och år kan
Fekalierna kan antingen samlas torrt eller separeras från spolvattnet efter en
kortare vattentransport i en separator. Denna utnyttjar centrifugalkraften,
gravitationen och ytspänningen för att separera partiklarna från vattnet. Med denna
typ av separation är det möjligt att separera mellan 58% och 85% av fekaliernas
växtnäring. Genom att återföra växtnäringen från urinen och fekalierna minskar
energianvändningen dels i avloppsreningsverket och dels genom att mineralgödsel
ersätts. Växtnäringen från urin och fekalier måste återföras till lantbruket för att
samhället skall kunna bli uthålligt.
För att recirkuleringen skall kunna anses vara säker avseende överföring av smitta
måste fekalierna hygieniseras innan användning. Om fekalierna samlas torrt är det
möjligt att hygienisera dem genom termofil kompostering. Bäst effekt fås om
fekalierna komposteras tillsammans med köksavfall. En beräkningsmetod för att
uppskatta de hygieniska säkerhetsmarginalerna utvecklades. Säkerhetsmarginalen
för hygienisering av fekalier tillsammans med köksavfall och gammal kompost
som strömedel var 37 gånger total inaktivering av enterovirus, den mest
temperaturtåliga organismen av de som utvärderades.
En annan metod för hygienisering av fekalier som undersöktes var kemisk
hygienisering med urea och med perättiksyra. Den mycket reaktiva perättiksyran
inaktiverade undersökta organismer inom 12 timmars behandling vid en dosering
mellan 0,5% och 1,0%. Den höga andelen torrsubstans (10%) i materialet ökade
kemikaliebehovet. Lägre torrsubstanshalt skulle signifikant minska doseringen för
att uppnå önskad hygienisering.
Hygieniseringen genom ureatillsats visade sig vara en lovande behandling. Genom
tillsats av 30 g ureakväve per kg material inaktiverades de undersökta
organismerna E. coli, Salmonella spp, Enterococcus spp, Salmonella typhimurium
28B fag och Ascaris suum ägg inom 50 dagars behandling. Sporformande
Clostridium spp var resistenta mot denna behandling. Genom att urean måste
brytas ned till ammoniak efter tillsats kommer behandlingen att ha en initial
fördröjning. Genom 2 månaders lagring av materialet, efter ureatillsats, erhålls ett
hygieniskt säkert gödselmedel. Dessutom ökar det tillsatta kvävet materialets
Ett urinsorterande och fekalieseparerande avloppssystem med hygienisering
genom ureabehandling verkar vara en intressant väg för att minska
resursförbrukningen och möjligen öka avloppssystemets hygieniska standard.
Den mätta dagen, den är aldrig störst.
Den bästa dagen är en dag av törst.
Karin Boye
List of papers
This thesis is based upon the following papers, referred to in the text by their
Roman numerals. Published papers are appended and reproduced with kind
permission of the publishers.
I. Vinnerås, B., Palmquist, H., Balmér, P., Weglin, J. Jensen, A., Andersson,
Å. & Jönsson, H. The characteristics of household wastewater and
biodegradable solid waste – a proposal for new Swedish norms. Submitted
to Urban Water.
II. Vinnerås, B. & Jönsson, H. (2002). Faecal separation for nutrient
management – evaluation of different separation techniques. Urban Water
in press.
III. Vinnerås, B. & Jönsson, H. (2002). The potential of faecal separation and
urine diversion to recycle plant nutrients in household wastewater.
Bioresource Technology 84:3, 275-282
IV. Vinnerås, B. Björklund, A. & Jönsson, H. (2002). Thermal composting of
faecal matter as treatment and possible disinfection method – Laboratory-
scale and Pilot-scale studies. Bioresource Technology in press.
V. Vinnerås, B. Holmqvist, A. Bagge, E. Albihn, A. & Jönsson, H. Potential of
disinfection of separated faecal matter by Urea and by PAA for hygienic
nutrient recycling. (Submitted to Bioresource Technology)
Notes on the authorship of the papers:
In Paper I, the investigation at Ekoporten was performed by Vinnerås and Weglin,
the calculation of the greywater composition by Balmér and the investigation at
Gebers by Andersson, Jensen, Jönsson and Palmquist. The proposed new
designing values was developed by Vinnerås, Palmquist, Jönsson and Balmér.
In Paper II, the planning of the investigation was performed by Vinnerås and
Jönsson. Vinnerås performed the investigation, the interpretation and analysis of
the results and the writing, with revisions by Jönsson.
In Paper III, the planning of the investigation was performed by Vinnerås and
Jönsson, and the sampling by Vinnerås and Weglin, Vinnerås carried out the
interpretation and analysis of the results and the writing, with revisions by
In Paper IV, the planning of the investigation was performed by Vinnerås,
Björklund and Jönsson. Björklund and Vinnerås performed the investigation.
Vinnerås carried out the interpretation and analysis of the results and the writing,
with revisions by Jönsson.
In Paper V, the planning of the investigation was performed by Vinnerås, Jönsson
and Albihn; Vinnerås and Holmqvist performed the investigation. The
interpretation of the results and the writing were carried out by Vinnerås, with
revisions by the other co-authors.
Introduction 13
Objectives 13
Background 14
Sewage treatment 19
Conventional large-scale central sewage treatment in Sweden 19
Conventional small-scale sewage treatment in Sweden 20
Recycling sewage systems 20
Urine-Diversion 20
Faecal separation 22
Blackwater diversion 23
Disinfection of sewage products 23
Storage 24
Thermal treatment 25
Chemical treatments 30
Environmental effects from sewage treatment 34
Composition of urine, faeces and greywater (Paper I) 35
Urine 36
Faeces 38
Greywater 40
Proposed new designing values for the Swedish wastewater fractions 42
Faecal Separation (Paper II, III) 42
Laboratory-scale evaluation of different separation techniques (Paper II) 45
Controlled pilot-scale separation of faeces and toilet paper by using whirlpool,
surface tension separation 46
Full-scale separation of faeces in the Ekoporten block of flats (Paper III) 50
Important issues when using faecal separation 51
Disinfection of faeces (Papers IV, V) 53
Thermal composting of separated faecal matter (Paper IV) 53
Disinfection of faeces 58
Chemical treatment of separated faecal matter (Paper V) 60
Treatment with peracetic acid (PAA) 60
Treatment with Urea 62
Risk of regrowth after disinfection treatments 63
Comparison of sewage treatment strategies 64
Comparison of different primary treatments 65
Nutrient recycling 66
Quality of the recycled products 67
Water emissions 68
Energy usage 68
Comparison of secondary treatments to attaining high hygiene standards 70
Concluding summary 73
Further research needed 77
References 78
Acknowledgements 87
The 90 naturally occurring elements have been present in some form since the
earth was formed, either in their ground state or combined in molecules. During
the evolution of life, biological systems have concentrated some of the elements
while others have been rejected, as a natural selection of the elements. Today,
approximately eleven elements contribute 99.9% of the total biomass. In the
human body, four elements contribute 99% of the total wet mass (H=62.8%,
O=25.4%, C=9.4% and N=1.4%). The other 0.9% are Na, K, Ca, Mg, P, S and Cl.
Seventeen other elements have been identified as being used by biological life
(Fraústo da Silva & Williams, 1997; Williams & Fraústo da Silva, 2000).
The remainder of the elements in the periodic table are in most cases used in some
type of commercial process in the current industrialised world. Therefore, they will
be found in the products of society. Apart from the elements, approximately
30 000 compounds are used as everyday chemicals and will probably be found in
the wastewater (Palmquist, 2001). However, in this thesis the emphasis is on the
elements, especially those essential to biological life.
The ideal situation in waste management would be that the bio-useful elements
would end up in one place for recycling into food, while the rest of the elements
and substances would be recycled elsewhere. However this is far from being the
case today, especially in the wastewater systems, where a lot of these elements and
substances are mixed. Both essential and lethal substances plus pathogenic
organisms from households and from industries are collected together and treated
in some way so as to minimize their immediate influence on society.
The main proportion of the nutrients in household wastewater is excreted in the
urine and the faeces, while only small amounts of toxic heavy metals and organic
pollutants are found in these fractions. The main pollutants are generally found in
the greywater from households (Paper I; NV, 1995) and in wastewater from
industries (Balmér, 2001). By collecting the urine and the faeces separately, it is
possible to collect a large portion of the nutrients in two relatively unpolluted low-
volume fractions.
The diverted urine is easily disinfected by storage (Höglund, 2001) while the
faeces, which contain the main proportion of the potential pathogens, have to be
actively disinfected. Today no simple, reliable, cost-effective and scale-
independent sanitation method is available for treatment of faecal matter.
The main objective of this thesis was to find a hygienic and environmental friendly
way to recover the nutrients from household wastewater and safely recycle them
back to arable land.
To achieve this, the composition of biodegradable waste and wastewater had to be
determined (Paper I).
Urine has been investigated in detail earlier and shown to be the cleanest and the
most nutrient-containing fraction; the methods for recovery of this fraction have
also been exhaustively investigated. However, the second most nutrient-containing
fraction, the faeces, has been the subject of very little study. Therefore, a sub-
objective of this study was to find easy ways to recover faecal nutrients while still
using the facility of flush toilets (Paper II, III).
A lot of pathogens can be found in the faecal fraction and before the faecal
nutrients can be utilized, they have to be disinfected in some way. Different
methods for disinfection were investigated, i.e. thermal composting (Paper IV) and
chemical treatment by urea and by peracetic acid (Paper V).
As an effect of the urbanisation and the increase in the population during the 19th
century, problems with transmission of diseases between the increasing numbers
of humans have occurred. One factor identified in such problems was the handling
of human waste. To solve this, water-borne sewage systems were built for the
urine, faeces and greywater.
The effect of the introduction of the sewage systems was a decrease in diseases
linked to latrines. Instead, new problems were introduced in the receiving waters.
Larger particles were accumulated in the area around the outlets, but by using
mechanical treatment before discharging the water this problem was removed.
The next problem identified was the discharge of oxygen-consuming organic
matter, which led to oxygen deficiency in the recipient waters. The absence of
oxygen led to death of fish and production of H2S and other unpleasant gases. By
introducing aeration of the wastewater and developing the technique of active
sludge, this problem was also solved.
In Sweden, the next problem identified was a secondary production of oxygen-
consuming organic material, mainly caused by phosphorus in the effluent. Within
a short period of time, this phosphorus was taken up by algae, allowing them to
produce more organic material. One gram of phosphorus corresponds to a
consumption of 140 grams of oxygen as the alga degrades after death. By addition
of metal ions, such as iron, aluminium and calcium, it was possible to precipitate
the phosphorus from the water into a more solid sludge (Tchobanoglous & Burton,
Nitrogen is another element that has caused problems in recipient waters, both as a
biochemical oxygen consumer during nitrification of ammonia to nitrate and, when
nitrogen is deficient in the surroundings, causing production of biomass. One gram
of nitrogen, as ammonia or nitrate, corresponds to a consumption of approximately
20 grams of oxygen as COD (Nord, 1995). Surplus of oxygen results in
nitrification and if followed by anoxic conditions and with organic carbon present,
the nitrate is reduced into gaseous dinitrogen, as is present in air (Tchobanoglous
& Burton, 1991). These two natural biochemical processes are also used in sewage
treatment plants to remove nitrogen.
The initial reason for establishing sewage systems, to improve sanitary standards
and decrease the transmission of diseases, had a positive effect towards diseases
within the cities. The pathogens was transported out of the cities and diluted in the
water recipients, where it sometimes can be found in a wider area both in surface
and ground water.
One of the main opponents to the introduction of water-borne sewage systems was
the agricultural sector. Farmers feared that they would lose a valuable fertiliser as
the latrines disappeared. This opposition weakened with the development of the
Haber-Bosch process, making it possible to transform dinitrogen from air into
ammonia and thereafter into nitrates by the Ostwald process during the early 20th
century. For the two other nutrients in focus, phosphorus and potassium, methods
for mining were developed and the need for organic fertilisers such as latrines
During the second half of the 20th Century, the focus on the use of natural
resources increased. The natural resources of fossil fuel (oil and gas) are predicted
to last for another 30 to 50 years. This will have a big impact on the production of
nitrogen fertilisers, which are heavily dependent on fossil fuel for the reduction of
dinitrogen to ammonia (Greenwood & Earnshaw, 1998). The largest single energy
requirement in the conventional production of rapeseed in Sweden is the
production of the mineral nitrogen fertiliser used (Hovelius & Hansson, 1999).
The major part of nutrients on farms is recirculated within the farm, by circulation
of materials such as roots and straw left on the field at harvest and by circulation
of manure as fertiliser (Steineck et al., 2000). However, there is still an outflow of
nutrients following food produced for human consumption (Figure 1).
Today, this outflow into society is mainly a one-way flow. To have a sustainable
society, the nutrients from farms have to be recycled back to the farms from
Figure 1 The flow of nutrients on the farm and into society.
In a healthy adult, the amounts of nutrients are in equilibrium within the body.
Thus all the plant nutrients consumed are excreted; normally via the urine or via
the faeces (Guyton, 1992), although some metals are also found in the sweat
(Schroeder & Nason, 1971).
Due to the origins of urine, faeces and greywater, there are major differences in
their composition. Urine is composed of material that is metabolised in the body
and extracted from the bloodstream into the urine by the kidneys (Guyton, 1992).
The faeces are composed of both material extracted from the bloodstream and
material that just passes through the intestine undigested. In the faecal fraction of
the wastewater in the Western world, there is also a quantity of toilet paper. The
greywater is a waste fraction consisting of water used by the households and some
liquid food waste.
The composition of the urine reflects the consumed material, especially the parts
that are digested and taken up by the intestines. The main component in urine is
water but also excreted in the urine are the main proportion of the household
nutrients (Figure 2).
Nutrient flow (g/p,y)
Biod. Solid
Figure 2. The content of nitrogen (N), phosphorus (P) and potassium (K) in the fractions of
biodegradable solid waste and wastewater from households (Paper 1).
The nutrients in urine are in a water-soluble form. Nitrogen is mainly found as
urea (80%), ammonia (7%), and creatine (6%), while the remainder is mainly free
amino acids or shorter peptides (Johnston & McMillan, 1952; Lentner et al., 1981;
Guyton, 1992; Kirchmann & Pettersson, 1995; NV, 1995; Fittschen & Hermann,
1998; Almeida et al., 1999). This, combined with the biochemical activity of
urease (Alef & Nannipieri, 1995) transforming the urea into ammonia and carbon
dioxide (Vinnerås et al., 1999), indicates that the nutrients in urine are readily
plant available.
In the urine, the phosphorus is mainly found as inorganic phosphates (>95%) and
the potassium mainly as free ions (Berger, 1960; Lentner et al., 1981; Guyton,
Approximately 50% of the faecal nitrogen is water-soluble (Trémolières et al.,
1961). Of this, 20% is ammonia, biochemically degraded from urea, peptides and
amino acids. About 17% of the total nitrogen content is found in living bacteria
and the remainder is mainly found as organic nitrogen combined in molecules such
as uric acid and enzymes (Lentner et al., 1981).
The main proportion of the phosphorus in the faeces is found as undigested
mineral calcium phosphates. Potassium on the other hand is mainly found in its
ionic form in equilibrium with the liquids outside the intestine (Berger, 1960;
Guyton, 1992; Fraústo da Silva & Williams, 1997).
The low amount of heavy metals (metals with a density over 4.5 g cm-3) in the
urine is mainly due to the minor biological usage of heavy metals regulating their
uptake and thereby their excretion. However, small amounts of nonessential heavy
metals are taken up, in some cases in relatively large amounts. This especially
applies to metals similar in size and charge to those used in biochemical reactions.
This is the case with the uptake of cadmium for example, which uses iron-
pathways. The dose and whether the metals are eaten or inhaled also influences the
uptake, where the uptake of inhaled metals is much larger (Kehoe et al., 1940;
Joselov et al., 1967; Vahter et al., 1991; WHO, 1991, 1992, 1995; Kim &
Fergusson, 1993; Engqvist, 1998).
Due to the low uptake of heavy metals, the majority of the metals consumed just
pass through the body and are therefore found in the faeces. However, the
contribution of heavy metals to sewage water from both the urine and faeces is
considerably lower than the amounts coming from the greywater (Figure 3).
Cu Cr Ni Zn Pb
Heavy metal flow (mg/p,y)
Biod. Solid
Figure 3. The amounts of the heavy metals Cu, Cr, Zn, Ni and Pb in the fractions of
biodegradable solid waste and wastewater.
During the past few decades in Sweden, the focus has been on excluding mercury
and cadmium from society, e.g. mercury-containing thermometers have been
collected and exchanged for digital ones, and the flow of these two metals has
decreased (Lohm et al., 1997; Balmér, 2001). The use of mercury in dental fillings
results in a significant load of mercury to the faeces (Engqvist, 1998) (Figure 4).
Cd Hg
Heavy metal flow (mg/p,y)
Biod. Solid
Figure 4 The amounts of the heavy metals Cd and Hg in the fractions of biodegradable solid
waste and wastewater.
While the greywater contains the major proportion of the heavy metals (Figure 3;
4) it contains only a minor proportion of the nutrients (Figure 2). The nutrients in
the greywater are mainly inorganic and the amounts depend on the amounts used
in households. Phosphorus and potassium are used in detergents and their usage is
reflected in the amounts of these elements in the greywater.
The heavy metals in the greywater have different origins. One main source is dust,
especially small particles impossible to vacuum, which have to be wiped off
(Månsson, 1992). This dust content in domestic cleaning water is a major source of
heavy metals (Moriyama et al., 1989; Kim & Fergusson, 1993; Comber & Gunn,
1996; Koch & Rotard, 2000).
Some heavy metals can also be found in chemicals used in households, e.g.
cadmium in phosphorus-containing detergents and zinc in anti-fungal shampoo
(dandruff shampoo). This kind of usage is reflected in the metal content in the
The total volume of the fractions also differs greatly. The greywater is a high
volume fraction and in Sweden it contributes between 66 and 75% of the total
volume of household wastewater. This corresponds to 24 to 65 m3 person-1 (p-1)
year-1 (y-1) (Palmquist, 2001). In systems with a flush toilet, the second largest
volume consists of flushwater. Today, the volume of water used per flush normally
varies from 2 to 6 litres per flush. The total annual volume, if 6 litres are used for
each flush, corresponds to 18 m3 p-1 y-1.
The fractions flushed, urine and faeces, are very low volume compared to the
water used. The collected urine volume corresponds to 365-550 litres p-1 y
-1 and
the faeces, toilet paper included, corresponds to approximately 40 kg wet weight
per person and year (Paper I; Lentner et al., 1981; NV, 1995; Hellström &
Kärrman, 1996; Hanaeus et al., 1997; Jönsson et al., 1997, 1998, 2000; Höglund et
al., 1999; Otterpohl et al., 1999).
As we can see, the nutrients leaking from urban society are mainly found in the
fractions of urine, faeces and biodegradable solid waste. If we were able to recycle
these three fractions to agriculture, the need for extra inputs of nutrients to farms
would decrease and the consumption of fossil resources would also decrease,
making it possible to increase the sustainability of society.
Sewage treatment
The sewage treatment of today in Sweden is normally performed either on a large-
scale at central treatment plants or on a smaller local-scale even down to detached
households using cleaning processes such as soil filters.
In a more global perspective, the most common way is to not treat excrement or
wastewater at all. In the absence of toilets, people defecate in fields or in some
cases in holes that are dug each time. The greywater produced is either wasted
outside the building or used to water plants in the surroundings.
In a lot of places with wastewater systems, the untreated wastewater is just
dumped into a recipient water body nearby. This gives problems both with
eutrophication and with transmission of pathogens and a valuable nutrient source
is also lost as a pollutant, which leads to the possible loss of the water body as a
resource for drinking water and for fish production.
Conventional large-scale central sewage treatment in Sweden
The majority of the household wastewater in Sweden today is collected and led to
a central sewage treatment plant where it is treated mechanically combined with
biological and/or chemical treatment.
Larger particles are mechanically removed when the water passes through sieves
on entering the treatment plant. Biological and chemical treatment then removes
the majority of the BOD (biological oxygen demand), phosphorus and, in larger
plants in Southern Sweden, also nitrogen.
The BOD is mainly removed by biological oxidation of organic compounds,
resulting in gaseous carbon dioxide. By having alternating zones with aerobic and
anoxic treatment, it is possible to first nitrify the ammonia nitrogen and then
denitrify the nitrates. This process needs a lot of space and energy for aeration and
the result is a transformation of the plant available nitrogen into gaseous
dinitrogen. To make that airborne nitrogen plant available again, a lot of energy,
mainly fossil oil or gas, is used in the Haber-Bosch process to produce ammonia
from dinitrogen (Greenwood & Earnshaw, 1998).
Phosphorus is in Sweden normally removed from sewage water by chemical
precipitation. This can be performed throughout the process and the methods used
for precipitation are mainly addition of iron (Fe2+ or Fe3+), aluminium (Al3+) or
calcium (Ca2+) ions. With the precipitated metal phosphates, some organic matter
and often a lot of heavy metals follow from the wastewater into the sludge
(Balmér, 2001). Biological methods for phosphorus removal are available today,
but they are not common in large-scale systems in Sweden. Other incoming
nutrients such as potassium more or less just pass through the treatment plant
unaffected (Balmér et al., 2002).
Conventional small-scale sewage treatment in Sweden
About half of the emissions of phosphorus from sewage to recipient waters in
Sweden are caused by about one million houses, both permanent and summer
houses, not connected to a central treatment plant.
The wastewater from these systems is mainly treated in small local systems
consisting of a septic tank and a soil filter or an infiltration unit. The BOD removal
of these systems is generally good but the removal of nutrients is often little to
medium (Aaltonen & Andersson, 1995). The infiltration systems are also hard to
supervise regarding leakage to the groundwater, thus undiscovered contamination
both by nutrients and pathogens can occur.
The potential for recycling the nutrients is small, since the only possible fraction is
the sludge from the septic tank. However, its nutrient content is low and its heavy
metal content high (Almedal, 1998; Svensson & Mattson, 1999).
Recycling sewage systems
To decrease the discharge of eutrophying substances to recipient waters and to
decrease the need for fossil resources in agriculture, the two most nutrient
containing fractions, the urine and the faeces, can be collected separately and
recycled to agricultural production.
The collection of the toilet fractions can either be performed separately, where the
urine is diverted and the faeces are collected dry or separated from the flushwater
after a short transport, or by blackwater collection, where both the urine and faeces
are collected as one fraction.
The greywater still has to be treated, but this treatment can then be more adapted
to the special conditions required by the greywater. However, in this thesis the
treatment of greywater is not considered.
One of the major disadvantages with the current sewage system in Sweden is the
problem of efficiently recycling the nutrients, in an uncontaminated way.
By collecting the urine and the faecal fractions separately from the rest of the
sewage water, the majority of the nutrients (Figure 2) are contained in a small
fraction relatively unpolluted by heavy metals (Figure 3; 4).
By using urine-diverting toilets (Figure 5), it is possible to collect the most
nutrient-containing fraction in a concentrated way. This kind of toilet has two
separate bowls where the smaller front bowl collects the urine, which in some
cases is flushed with a small amount of flushwater. In the rear bowl the faeces and
used toilet paper are collected. The rear bowl can either be flushing or non-
Figure 5. Two double flush urine-diverting toilets, Dubbletten and Gustafsberg, and one
single flush toilet where the faeces are collected dry, Wost Man Ekology ES.
If the urine is collected separately, the main proportion of the nutrients from the
household is collected in a small fraction with a low heavy metal content. Due to
the rapid degradation of urea into ammonia and carbon dioxide, the pH of the urine
mixture rises from approximately 7 (Lentner et al., 1981) to above 8.5 (Vinnerås et
al., 1999; Jönsson et al., 2000; Höglund et al., 2000).
The high ammonia content of the collected urine mixture makes it very corrosive
towards metals and it also increases the ageing of PVC (Greenwood & Earnshaw,
1998). The corrosion of metals means that if metal pipes are used, metals pollute
the initially low heavy metal-containing urine. This was identified as a problem in
households that had urine-diverting toilets with half a metre of copper pipe
connecting the urine bowl and the soil pipes, which increased the copper content
of the urine mixture by more than 50 times (Jönsson et al., 1997, 1998, 2000;
Vinnerås et al., 1999).
If the ammonia content is over 1 mg l-1, the pH is over 8.8 and no fresh urine is
added, storage of the urine mixture for one to six months, depending on the
temperature, inactivates any non spore-forming pathogens present, so the urine can
then be recycled as a fertiliser to agriculture with negligible hygienic risks
(Jönsson et al., 1997, 2000; Höglund et al., 1998; Höglund, 2001).
If the urine is collected and uncontaminated by heavy metals, an unpolluted and
potent organic fertilizer that contains mineral nutrients is obtained. Field trials and
pot experiments have shown that diverted human urine is comparable to mineral
fertilisers. For nitrogen, the fertilising effect is comparable to, or just a little bit
poorer than, mineral fertilisers while for phosphorus, the fertilising effect is
comparable to, or a little bit better than, mineral fertilisers (Kirchmann &
Pettersson, 1995; Elmqvist et al., 1998; Johansson et al., 2000).
Faecal separation
The second largest contributor of nutrients to wastewater is the faeces (Paper I;
Figure 2). This fraction is also the smallest volume fraction in the household
wastewater (NV, 1995; Kärrman et al. 1999; Paper I). The faeces can either be
collected dry beneath the toilet or separated from the flushwater after a short
transport (Paper II, III).
If using a single flush toilet (Figure 5) or a non-flush dry toilet or squatting plate
(Figure 6), the faecal matter and the toilet paper (if used) are collected in a
container or vault directly beneath the toilet. When the urine is diverted, problems
with smell and flies are significantly reduced (Fittschen & Niemczynowicz, 1997;
Del Porto & Steinfeld, 1999).
Figure 6. A non-flush dry urine-diverting toilet and a urine-diverting squatting plate
Commercially available today for separation of faecal matter from flushwater are
two systems, one that has a filter that collects the particles from the flushwater
over several months, building up a filter cake (Otterpohl, 2001; Hellström &
Johansson, 2001). The other system separates the faecal particles in a whirlpool,
surface tension separator, by using a combination of whirlpool effect, gravitation
and surface tension (SP, 1992; Del Porto & Steinfeld, 1999; Paper II, III).
By using this kind of wet separating system, the collection of the faecal matter
does not have to be performed directly beneath the toilet, but can instead be done
at another location, for example in the basement of a multi-storey building. Thus
the comfort of having a flush toilet in the house is still possible even when
recycling the nutrients from the urine and faeces using a source separating sewage
Blackwater diversion
The urine and faeces can be collected together, but separately from the greywater.
If this is performed dry, in bucket latrines or pit latrines, there is an impending risk
of problems with both smell and flies.
If the blackwater is collected in a flushed system, only small volumes of
flushwater can be used, to keep the volume down and the dry matter content up.
Normally these systems use low-pressure sewers, using vacuum to transport the
material from the toilet to a collection tank (Skjelhaugen, 1999a).
The advantage with this kind of system is that all of the excreted nutrients are
collected both from the urine and the faeces, unlike urine-diversion systems, where
some of the urine is lost (Jönsson et al., 1997, 1999, 2000; Höglund et al., 1999),
and faecal separation, where some of the nutrients are lost by extraction (Paper II,
The disadvantage with blackwater diversion is the need for highly developed
technology based on a vacuum system to keep the flushwater volumes low.
However, too small volumes can lead to blockages in the system, as small water
volumes sometimes will not manage to transport the particles to the collection
tank. This system has low robustness and is thereby sensitive to disruption and if
the system goes down, e.g. through blockage or leakage, the toilets do not work.
Disinfection of sewage products
Handling human waste has been a sanitary problem for a long time. This problem
is so old and large in scale that behaviour concerning excretion is regulated in
some religions and beliefs.
As mentioned previously, hygiene was one of the major reasons for introducing
wastewater systems into cities, but all hygiene problems were not solved by the
introduction of the water-borne system. The pathogens can in the water recipient
be transported and cause diseases somewhere else. Also the rest products if aiming
for recycling can contain sever amounts of pathogens.
When considering the recycling of sewage products to arable land, transmission of
diseases is one of the major issues. The Swedish farmers’ association (LRF) has as
one of the conditions for recycling sewage sludge set a restriction of no viable
Salmonella bacteria in the sludge. In a study performed by Albihn et al. (2001)
more than 70% of the sludges analysed contained Salmonella, even sludge after
mesophilic (30-45°C) anaerobic digestion.
Pathogens are also found in products from source-separating systems. Therefore,
all products from the wastewater system have to be disinfected before they can be
considered for use as fertilisers. Treatment methods discussed in this thesis are
storage, thermal treatment and chemical treatment.
Storage is a low impact alternative for attaining improved hygiene status of the
material intended to be recycled. The disinfection effects of storage depend on the
stored material. As a function of time, some inactivation of pathogens occurs
during storage (Brock et al., 1994). However, storage is not a reliable method, and
some of the pathogenic microorganisms have a dormant state in which their
tolerance towards the surrounding environment is higher than in their vegetative
The efficiency of storage on pathogen control is higher at higher temperatures
(Feachem et al., 1983; Mitscherlich & Marth, 1984). In the Swedish climate where
the temperature is around and below zero for most of the winter, the storage has to
be performed partly during the warmer half of the year.
Storage of organic waste products can also produce other problems, especially
towards the surrounding environment through emissions, both of eutrophying,
acidifying and greenhouse gases (Flodman, 2002).
Storage of human urine
Due to the degradation of the urea in urine to ammonia and carbon dioxide, as
early as in the soil pipes (Höglund et al., 1999; Vinnerås et al., 1999; Jönsson et
al., 2000), the environment in the urine mixture becomes toxic towards most of the
microorganisms present.
Höglund et al. (1998) have investigated the survival of bacteria in human urine at
different water-diluted urine mixtures, i.e. undiluted, 1:1 and 1:9 urine to water.
Aeromonas hydrophila, Eschericha coli, Pseudomonas aeruginosa, two
Salmonella sub-species and faecal Streptococci were studied and all of them had a
D value (decimal reduction value, i.e. 90%) of <35 days in the 1:9 diluted urine at
a storage temperature of 4°C. As expected, significantly shorter D values were
detected for more concentrated urine mixtures and at higher storage temperature
(20°C) (Höglund, 2001).
Clostridium perfringens and the Salmonella typhimurium 28B phage did not show
any reduction in viability compared to the control (Höglund et al., 1998). A more
recent study of the Salmonella phage and of Rhesus Rotavirus showed no
significant reduction at 5°C but at 20°C the D values were 71 and 35 days,
respectively (Höglund et al., 2002). Ascaris suum ova were also stored in the
different urine mixtures and a reduction in viable organisms of 15-20% was
achieved during 21 days of storage at 20°C (Höglund et al., 1998).
A longer study of the survival during storage of Cryptosporidium parvum in urine
has also been performed (Höglund & Stenström, 1999). In that study, the survival
was monitored during 133 days at pH 5, 7, and 9 of the urine mixture. At pH 9 the
D value was 29 days and at pH 7 the D value was 207 days, which was 50% longer
than in one of the controls.
These studies (Höglund et al., 1998; Höglund & Stenström, 1999) indicate that the
main effect on inactivation of microorganisms from storage of human urine is a
combination of temperature, ammonia concentration and pH, similar to the results
from ammonia treatment of other sewage products (Burge et al., 1983; Cramer et
al., 1983; Werkerle & Albrecht, 1983; Allievi et al., 1994; Gaspard et al., 1995;
Gantzer et al., 2001).
A recommendation for urine storage to attain acceptable safety limits has been
developed for Swedish conditions and use of the urine as a fertiliser after storage
at different temperatures, to different crops and different uses of the crop produced
(Jönsson et al., 2000; Höglund, 2001). These recommendations vary from shorter
storage times (1 month) at 4°C where the urine can be used on crops that are
processed before being used as fodder or food, to longer storage times (6 months)
at 20°C where the urine can be used on all kinds of crops, even those that are
consumed raw by humans.
Storage of other wastewater products
The inactivation rate of microorganisms during storage of faeces and sludge rich in
organic matter and poor in free ammonia is lower compared to that in materials
that have higher amounts of ammonia.
The mechanisms dominating the reduction of pathogens during storage of
wastewater products are mainly competition from other organisms and starvation.
Due to this non-specific effect that differs according to local circumstances, the
effects of storage are difficult to predict and to specify.
Tests on the survival of different Salmonella sub-species during storage of faeces
and sewage sludge have shown that the materials still contain viable organisms
after 100 days of storage (Mitscherlich & Marth, 1984).
When storing sludge with high water contents, fermentation of the sludge occurs
and produces fatty acids that lower the pH (Haug, 1993; Skjelhaugen, 1999a). This
drop in pH can result in a higher inactivation rate of pathogens (Mitscherlich &
Marth, 1984; Brock et al., 1994). However, the anaerobic degradation of the
sludge also produces gases such as methane and nitrous oxide (Flodman, 2002),
which both are very potent greenhouse gases.
Thermal treatment
Sensitivity towards heat varies between organisms. Some organisms can stand a
higher temperature for a longer time than others, while some other organisms can
stand a lower temperature for a longer time. This is exemplified by the
Enteroviruses, which are more resistant to high temperatures than Ascaris, but the
latter are more resistant at lower temperatures (Figure 7).
45 50 55 60 65
T (°C)
t (h)
Taenia eggs
Ent. Hystolica
Figure 7. The time for total inactivation of nine pathogens at temperatures between 45°C
and 65°C. The equations for the deactivation times are taken from Feachem et al. (1983)
(Paper IV).
It is also possible to present the deactivation time for these organisms as a function
of the temperature (Table 1). For Ascaris the inactivation can only be guaranteed
above 45°C and for some other organisms the function of inactivation is changed
at temperatures below 45°C (Feachem et al., 1983). Therefore, these equations are
only functional above that temperature.
Table 1. Equations derived from Feachem et al. (1983) for the time in hours (t) required to
attain no viable organisms of different pathogens at different temperatures (T) in Celsius
(Paper IV)
Organism Type Equation (Ed)
Enteroviruses Virus
Cholera Bacteria t=0.89×10-0.2117(T-45)
Salmonella Bacteria t=75.4×10-0.1466(T-45)
Shigella Bacteria t=13.8×10-0.1369(T-45)
Enteric Hystolica Protozoa t=21.3×10-0.2806(T-45)
Ancylostoma Helminth t=9.31×10-0.1340(T-45)
Ascaris Helminth t=177×10-0.1944(T-45)
Schistosoma Helminth t=10.0×10-0.1844(T-45)
Taenia Helminth t=6.6×10-0.1306(T-45)
The inactivation of the pathogens is logarithmic and even if the time interval is
shorter than that needed to achieve no viable organisms, a partial inactivation is
obtained, which can be seen as a number of logarithms of the total inactivation
(Stanbury et al., 1995). The sum of these partial inactivations shows the number of
times we could expect to achieve the state of no viable organisms. From that value,
we have a quantified safety margin of how many times the inactivation was
obtained during the time of treatment.
)()()( 33
ddd TE
X++++=Σ (1)
ΣX= number of times total die-off is achieved;
Ed = the time for total die-off at the average temperature T during the time interval
t. Calculated from equations in Table 1.
t= time interval (h) for the temperature T
The safety margin (Table 1) was initially set by Feachem et al. (1983), and these
margins are based on several different measurements of survival as a function of
time and temperature. All of the compiled inactivation at different temperatures,
most of the time when no viable organisms were detected, was plotted in a time-
temperature diagram. From that plot, a safety margin was set for treatment time at
different temperatures where no viable organisms should be possible to detect, in
most of the cases the line was set well above the determined time and temperature
for no viable organisms in the laboratory studies (Feachem et al., 1983). To
specify a state of total die-off, the Parental drug association determined 12 log10
reduction of organisms as a level of sterility (Enzinger & Motola, 1981), which
was assumed to be equal to the level of total die-off according to Feachem et al.
A plot of the safety margins at different temperatures for several pathogens set by
Feachem et al. (1983) is presented in Figure 7. If this safety margin is reached, no
viable organisms should be detectable within the treated material. Using Equation
1 to calculate the theoretical number of times the state of no viable organisms is
achieved will give a relative estimation of how safe this kind of treatment can be.
Some bacteria can form spores, which are a dormant state where the bacteria are
very resistant to external stress, such as heat (Brock et al., 1994). Other pathogens
such as prions are also very stable towards heat. The moisture content of the
material affects the efficiency of the thermal inactivation (Stanbury et al., 1995;
Turner, 2002), as the heat transfer to the organisms and thus the inactivation is
more efficient when moisture is present.
To monitor the effects of temperature, indicator organisms can be used.
Alternatives here are either to use pathogenic microorganisms such as Salmonella,
Ascaris and Enteroviruses, or to use microorganisms with known deactivation
times at certain temperatures. For sterilisation of substrates, Bacillus
Stearothermophilus is one of the most thermo-resistant spores known and it is
normally used as an indicator of the efficiency of the sterilisation process
(Stanbury et al., 1995).
For safety reasons it would be preferable if all pathogens were killed. However, it
is not possible to secure total die-off, but only determine a state where no viable
organisms can be determined. If the conditions are changed, there is a risk for
regrowth of pathogenic bacteria, even if only one single bacterium survives.
Generating the heat for thermal treatment is mainly done in two different ways,
either an external source for heat, e.g. electricity or oil, or the organic matter in the
material, e.g. incineration or thermal composting. A combination of the two can
also occur, i.e. burning of biogas produced by anaerobic digestion of organic
Using an external heat source is expensive but the advantage is the possibility to
have a fully controlled process. In cases of further treatments not generating high
temperatures, e.g. mesophilic anaerobic digestion, or materials not containing
enough organic material for heat production, e.g. diverted urine, external heating
has to be used to achieve good hygiene standards by pasteurisation.
Another alternative for generating heat is to use the organic matter in the material
to produce the heat. One major advantage of this is that when the material is
degraded it is stabilised and the risk for regrowth then decreases (Sidhu et al.,
2001). The two most common treatment alternatives are incineration and thermal
If the material contains enough dry matter, incineration can be an alternative. By
this treatment, the material is stabilised, pathogens are inactivated and heat is
produced. However, before incineration the raw untreated matter has to be
transported to a treatment plant and to avoid negative environmental effects from
the incineration, the gases produced have to be cleaned. Some valuable nutrients,
namely nitrogen, are lost during the process.
Under aerobic conditions, microbiological digestion produces an excess of heat
energy. If that heat is captured by insulating the process, either with specialized
insulation material or a thick layer of organic matter, the temperature of the
material increases up to 80°C (Haug, 1993; Epstein, 1997). The composting can
either be dry; approximately 35-55% dry matter content, or liquid, 2-10% dry
matter content.
In liquid composting, the treatment has to be performed in a reactor with an active
oxygen supply. To get aerobic conditions, the aeration has to be performed in the
bottom part of the reactor, thereby requiring a lot of energy. The two most energy
using parts of the process are the aeration/mixing and the foam cutting
(Skjelhaugen, 1999b). Normally, aeration is also used for mixing or the mixer is
used to finely distribute the air in the reactor (Stanbury et al., 1995).
In a study by Skjelhaugen (1999b), two commercial aerobic reactors were
investigated (17 m3 and 32 m3). The small reactor consumed approximately 30
kWh m-3 and the larger approximately 20 kWh m-3 based on a 7 day retention time
When finely distributed bubbles were blown into slurry, foam was produced. To
damper this, a foam cutter was used, thereby stopping foam from overflowing the
reactor. The foam cutter used about 5-6 kWh m-3 (Skjelhaugen, 1999b).
The advantage of liquid composting is the use of a completely mixed reactor that
allows full control of the process and process temperature to assure thermal
disinfection. The outgoing air is also controlled since it passes a heat exchanger
that condenses outgoing moisture and ammonia, which are then refluxed into the
reactor, thereby capturing 50% of the ammonia during summer and 75% during
the rest of the year (Skjelhaugen, 1999b). The air is then led through a bio-filter
where most of the organic substances and other odour-producing substances are
captured and decayed, thereby avoiding problems with odours in the surroundings.
The high water content (91-98%) and the low degradation of the organic matter
(10-20%) (Skjelhaugen, 1999b) give large volumes to handle and distribute on
farms. The low degradation of the organic matter indicates a risk for anaerobic
activity during storage before use, which will have large negative impacts on the
environment from production of methane, a potent greenhouse gas.
An alternative to the technique of the intense reactor in liquid composting is dry
composting, where the preferable dry matter content is approximately 35% or
higher. By using amendments with very high structure, it is possible to have
somewhat lower dry matter content and still maintain an aerobic process (Haug,
1993; Epstein, 1997).
To estimate whether the energy in the material for composting is high enough for
thermal composting, Haug (1993) developed two rules of thumb. The highest
energy usage in dry composting is the evaporation of water. Therefore, it is easy to
use the ratio of mass of water to mass of biological volatile matter (W) to
determine if the energy is sufficient to evaporate the water. If W (Equation 2) is
below 8, the energy in the organic matter is enough to produce elevated
temperatures in the compost. If W is greater than 10, the energy is not enough to
evaporate the water in the material and therefore less drying of the material occurs
during the process (Haug, 1993) and no especially high temperatures are reached.
= (2)
mH2O= mass of water
mBVS= mass of the biological volatile substance
When using Equation 2, the assumption is that all of the organic matter has the
same energy content per unit mass. This is not true in practice and if the energy
content in the composting material is known, Equation 3 gives more detailed
information on whether the heat production of the material is high enough for
thermal composting. If the ratio of the heat released to the weight of water (E) is
about 2.9 MJ g-1 H
2O, the energy content is sufficient for composting and for
evaporating the water (Haug, 1993).
= (3)
Hr= heat released
mH2O= mass of water
To attain temperatures high enough for thermal inactivation of pathogens, the heat
has to be kept within the material. Some parts of the material still hold a lower
temperature, i.e. the surface of an open heap and the area around the incoming
ventilation air. Thus the material does not have a homogeneous temperature and
not all of the material is thermally disinfected, giving less inactivation of
pathogens in the material. To get all material treated at elevated temperatures, it
has to be turned so the parts that are in low temperature zones are moved into high
temperature zones. To estimate the amount of surviving microorganisms if there is
total inactivation in the high temperature zones and no inactivation in the low
temperature zones, Equation 4 can be used (Haug, 1993).
tfnn (4)
nt= number of surviving organisms
n0= number of organisms initially present
N= number of pile turnings
f1= fraction of compost material in low temperature zones
Equation 4 can only be used if there is no regrowth of microorganisms in the
material within the low temperature zones. Helminths and viruses cannot grow
within the compost, but bacteria can. However, the equation will still give good
information about the safety margins achieved from the treatment used.
Chemical treatments
Addition of chemicals to sewage products can be used for disinfection.
Traditionally this has been done by addition of ash to latrines (Del Porto &
Steinfeld, 1999). Different chemicals used have different effects; the main effects
are those from use of acids (phosphoric acid), bases (ammonia) and oxidising
agents (chlorine). The effects of these chemicals are mainly inactivation of protein
(surface or cellular) or inactivation of the DNA/RNA strand (Burge et al., 1983).
When choosing chemicals for use as disinfectants for sewage products, an
additional effect to take advantage of is the agronomic value of the substances in
the disinfectants, e.g. use of Ca(OH)2, NH3, KOH and PO4-3 (Allievi et al., 1994).
By using chemicals having an agronomic value for the disinfection, the treatment
is not merely an expense but also a way to increase the fertilising value of the
Chemical treatment with bases
When comparing the disinfection efficiency of ammonium hydroxide (NH4OH)
and potassium hydroxide (KOH) added in similar concentrations as base
equivalents, Allevi et al. (1994) found a much higher inactivation of bacteria by
the ammonia-based compound. This was probably due to the higher content of free
ammonia for the samples where ammonium was added.
To observe the effects of ammonia on viruses at different pH values, Cramer et al.
(1983) tested polioviruses and the bacteriophage F2 with a constant ammonia
concentration at different pH (7.1-8.6), thereby varying the concentration of
uncharged ammonia (NH3). At pH below 8 the inactivation was slow, D (the
Decimal reduction value) >70h for polioviruses and >200h for F2 viruses. Above
pH 8, the inactivation was more rapid but the F2 virus was found to be 4.5 times
more resistant towards ammonia than the poliovirus. Therefore, the bacteriophage
F2 was proposed for use as a deactivation indicator of Enteroviruses in sewage
effluents disinfected by ammonia (Cramer et al., 1983).
The amount of free ammonia depends on the temperature, the pH and the
ammonium concentration (Svensson, 1993). For deactivation using ammonia,
different pH and temperatures have been tested and for bacteria a pH over 10 has
proved to be sufficient for non-detection of viable organisms. According to Allevi
et al. (1994) temperatures above 10°C are sufficient for having good effects and at
temperatures below 5°C no effect from the ammonia will be detected. This is
similar to the results Höglund (2001) found in urine that was stored at different
temperatures, where no inactivation of Salmonella typhimurium 28b phage and
rotavirus occurred during 6 months of treatment at 5°C. However at a temperature
of 20°C, the D value for the same phage was 71 days, which was more than double
the time for the D value of Rotaviruses treated in the same way, D=35 days
(Höglund et al., 2002).
The inactivation mechanisms of viruses by ammonia have been investigated
(Burge et al., 1983; Turner & Burton, 1997), and for single strand RNA viruses
such as the poliovirus, the inactivation is caused by RNA rupture for temperatures
below 50°C. For double strand viruses of both RNA and DNA, which are more
stable, the deactivation is mainly caused by denaturation of the surface coating
proteins (Burge et al., 1983; Turner & Burton, 1997).
For inactivation of Ascaris eggs in sewage sludge, a treatment of 2 months with an
initial pH of 12.5 is required (Gaspard et al., 1995). In the same study, amounts of
added lime giving a lower pH combined with a shorter incubation time resulted in
a reduction of 66%. This is comparable to the results found by Carlander &
Westrell (1999) for addition of ash to faecal matter, where the inactivation after
nine weeks of storage was found to be between 50% and 100%.
Gantzer et al. (2001) investigated liming of sewage sludge reaching a pH of 12.4
for approximately 24h, without achieving a significant reduction in nematode
eggs. They identified the application of lime to the sludge as the probable reason
for the bad performance. This emphasises the problems with the application of
chemical disinfectants and the fact that one must ensure that the chemicals are
homogeneously mixed into the material. Ascaris eggs seem to be resistant to high
pH for short periods, at least when the pH is increased by addition of non-
ammonia containing bases (Gaspard et al., 1995; Carlander & Westrell, 1999;
Gantzer et al., 2001).
The advantage of ash or lime addition to fresh latrine waste is a decrease in both
smell and flies. One major disadvantage with the use of ash is the difference in
alkalinity depending on the wood used. Therefore, no firm recommendations can
be made regarding the use of ash in general.
One main effect from increasing the pH is the amount of free non-ionised
ammonia. To get free ammonia the pH has to exceed 8 (Turner & Burton, 1997).
Thus, when the pH is regulated by ammonia it does not have to be especially high,
as in diverted human urine the average pH is between 8.5 and 9 (Höglund et al.,
1998; Höglund & Stenström, 1999; Höglund, 2001) and addition of ammonia-
containing chemicals to sewage sludge to produce a pH of 10 is enough (Allievi et
al., 1994).
Also when using other bases to increase the pH, one main effect is from free
ammonia (Herniman et al., 1973; Burge et al., 1983; Allievi et al., 1994). When
using other bases the ammonia concentration is lower, and to reach a concentration
of free ammonia high enough for deactivation of pathogens the pH has to be
higher. For example, for addition of lime to sludge the pH has to be over 12.5 to
have effects similar to ammonia addition to a pH of 10 (Burge et al., 1983;
Werkerle & Albrecht, 1983; Allievi et al., 1994). At a pH over 11 will almost all
of the ammonia be found in the uncharged state and at pH of 12.5 will probably
the high pH itself also have effects towards the viability of the organisms.
The main reason for the disinfection effect by ammonia is that at high pH and
temperature, the ammonia is found in its uncharged state, i.e. NH3 (Ward, 1978;
Cramer et al., 1983; Burge et al., 1983). Ammonia is in equilibrium between the
uncharged state and the +I state (Equation 5).
[NH4+]aq[NH3] aq +[H+] aq (5)
The equilibrium constant (Ka) in Equation 5 is determined by Equation 6. The
equilibrium constant Ka has been determined empirically (Svensson, 1993)
(Equation 7).
= (6)
Ka =)09018,0
10 +T (7)
T=temperature in Kelvin.
On measurement of the ammonia content, both charged and uncharged ammonia
are detected [NH4]tot=[NH4+]+[NH3]. Using the measured concentration of
ammonia combined with Equation 6, it is possible to calculate the amount of
uncharged ammonia in the solution (Equation 8) (Svensson, 1993).
NH +
][ 4
3 (8)
Thereby is it possible to determine the concentration of uncharged ammonia by
measuring the pH, the temperature and the total ammonia content (Equation 8).
Using peracetic acid (PAA) for disinfection
Peracetic acid (PAA) is an unstable organic peracid that can explode at high
concentrations and its decomposition is catalysed by contact with metals. PAA
normally occurs as an aqueous quaternary mixture in equilibrium with acetic acid,
hydrogen peroxide and water (Equation 9) (Arturo-Schaan et al., 1996; Liberti et
al., 1999).
Acetic acid Peracetic acid
The biocidal action of peracetic acid is much stronger than the effects of hydrogen
peroxide or acetic acid, from which it is derived and to which it ultimately
decomposes (Equation 9; 10) (Liberti et al., 1999).
H2O2 H2O+½O2 (11)
The biocidal effect of a PAA-H2O2 mixture is stronger than the effect of the
substances by themselves, which indicates a synergistic effect (Alasari et al.,
1992). The effects of PAA on microorganisms are reported to be disruption of the
sulphydril (-SH) and sulphur (-S-S) bonds within enzymes and cell walls,
dislocating the chemo-osmotic function of membrane transport (Lefevre et al.,
1992). In nematodes and helminths, the disinfection is ascribed to denaturation of
proteins (Baldry et al., 1991).
PAA has lately been introduced as an alternative to chlorination of outgoing
wastewater (Baldry & French, 1989; Baldry et al., 1991; Arturo-Schaan et al.,
1996; Liberti et al., 1999) and of materials and surfaces in hospitals (Thamlikitkul
et al., 2001), and it has been shown to have good effects towards microorganisms
even at very low concentrations, 0.15 ppm (Collignarelli et al., 2000), but the
efficiency against both bacteria and viruses is lower compared to chlorination
(Baldry & French, 1989; Baldry et al., 1991; Collignarelli et al., 2000). The
efficiency against viruses and some parasites is lower than it is towards bacteria
(Liberti et al., 2000).
The concentration of PAA has to be increased when increasing the organic matter
content in the treated material to attain efficient inactivation of microorganisms
(Baldry & French, 1989; Collignarelli et al., 2000). This is due to a combination of
reactions between the peracetic acid and the organic matter and incorporation of
the microorganisms and the organic matter.
When using this kind of oxidising agent on organic matter, a risk for production of
foam occurs (JV, 1997; Turner & Burton, 1997). If not all bacteria are inactivated
the risk for regrowth is high, especially when considering that the waste product,
acetic acid, can be used as a substrate by the bacteria (Liberti et al., 2000). Thus,
the treatment of wastewater products should be performed as close in time as
possible to the application as fertiliser.
Due to the high reactivity of PAA and its relatively harmless waste products
(Equation 10), no or very small negative effects concerning mutagenicity or
toxicity in plants have been identified from material directly applied after
disinfection (Monarca et al., 2000). Thus, PAA is best used as a disinfectant just
before application of the sludge to land.
Environmental effects from sewage treatment
The environmental effects differ when different sewage treatment options are
compared. The treatment alternatives chosen for comparison are urine-diversion
combined with faecal separation and a large-scale sewage treatment plant.
To be able to compare different systems for wastewater treatment the
environmental effects have to be evaluated and compared to each other. Kärrman
& Jönsson (1999) presented a standard for choosing effects to compare the
different systems by normalising them according to the effects from the sewage
system compared to the total effects in Sweden. From this normalisation, it was
possible to divide the fractions into priority groups. Priority Group 1 comprised the
factors contributing >10% to the total Swedish environmental impact of resource
usage/environmental effect per person, and Priority Group 2 the factors
contributing 0.1%-10%.
Priority Group 1 (>10%)
Discharges of N to recipient waters
Discharges of Cd, Hg and Pb to recipient waters
Recycling of P and N
Flows of Cd, Cu, Hg and Pb to arable land
Priority Group 2 (0.1%-10%)
Discharges of P to recipient waters
Discharges of Cu to recipient waters
Energy use
Cd to landfill
Global warming
By using the factors in the priority groups, there is little risk for pseudo
optimisation of the system by optimising factors that are insignificant on a global
scale. It should be borne in mind that local factors can have a great influence on
the environmental analysis. Even if some effects are insignificant on a global
scale, they can have significant effects towards the local habitat such as poisoning
of a local recipient water body or increasing the local transport to an unacceptable
level. Therefore, even if using the priority set by Kärrman & Jönsson (1999), the
local habitat has to be borne in mind when deciding on comparison factors.
Composition of urine, faeces and greywater
(Paper I)
Household wastewater consists of three separate fractions; urine, faeces and
greywater. In conventional treatment of the wastewater, where all the fractions are
treated as one, one important consideration is the composition of the mixed
fraction ending up in the sewage treatment plant. However when aiming to recycle
the nutrient content in the wastewater, the composition of the fractions differs
widely and thus also the treatment needed and the possible use of the fractions.
To be able to decide how to use the fractions, their composition has to be known.
During the 20th Century, the composition of the urine and faecal fractions was
determined in medicine and, as mentioned above, the fractions reflect the food
consumption of individuals so the expected amounts in the fractions are set as an
average of what to expect from an average person. During the late 1990s in
Sweden, the increased interest for recycling of household nutrients resulted in a
new interest in the composition of the fractions and led to some measurements
being performed, first on urine (Jönsson et al., 1997, 2000; Höglund et al., 1999;
Vinnerås et al., 1999; Vinnerås, 2001; Papers I, III) followed by an interest in the
composition of the different wastewater fractions (Papers I, III).
During the second half of the 20th Century, there was some focus on the greywater
as a result of effects caused by the content of this fraction. Eutrophication of
recipient waters by phosphorus discharges originating from detergents was one of
the first noted effects of the composition of the greywater. During the late 1990s,
the heavy metal content in greywater, which contaminates sewage sludge, has also
come into focus (Moriyama et al., 1989; Koch & Rotard, 2000; Enskog Broman,
More analytical data, new ways to collect the material and more sensitive
analytical methods for detecting elements even at low concentrations have made it
possible to scrutinise the old designing values (NV, 1995) and propose new
designing values for the composition of the fractions in household wastewater
(Paper I).
The old designing values for the different fractions was derived in different ways,
the composition of the urine and faeces were set by using several studies of the
composition of the fractions combined with data on actual Swedish food
consumption during the early 1990s (NV, 1995). In that study, 10% of food
purchased was estimated to end up in the fraction of biodegradable solid waste,
which was confirmed in the Ekoporten study (Vinnerås, 2001; Paper I). This
makes these values solid and in the measurements performed in Paper I some
values differed significantly. This was due to unexpectedly high amounts of some
of the heavy metals analysed as a result of the influence of local collection strategy
on contamination levels, e.g. copper pipes contaminated the urine mixture in
Ekoporten and zinc coating of the faecal pipes in Gebers contaminated the faeces
(Paper I).
In the field studies in the two housing areas, some significant differences were
detected, both compared to each other and to the designing values (NV, 1995), in
the heavy metal content in the urine and the faeces. The higher amounts could not
be excluded from being contamination from the collection strategies used and
thereby did it not affect cutinisation present designing values.
The present designing values for the greywater is based on one short measurement
in an ecological village, corresponding to approximately 400 person days,
combined with flow analysis in a few other separated wastewater systems. The
results from the measurements presented in Paper I, corresponding to 3050 person
days, combined with the calculation of the composition of the sewage sludge
produced in large Swedish cities generated reliable results of high validity, making
it possible to scrutinise and propose changes to the present designing values.
When the urine volumes collected in the studies both in Ekoporten and in Gebers
were recalculated to the urine production per person and year, significantly higher
volumes of urine were collected compared to the old designing values. In
Ekoporten 554 kg p-1 y-1 of urine was collected and in Gebers 510 kg p-1 y
-1. In
several other studies the amount of collected urine has corresponded to
approximately 550 kg p-1y-1 (Hellström & Kärrman, 1996; Jönsson et al., 1997,
1998, 2000; Vinnerås, 1998). Therefore a change in the present designing values
from 365 kg p-1 y-1 to 550 kg p-1 y-1 was proposed.
Due to some methodological problems when collecting the urine, both in
Ekoporten and Gebers, some misleading results were obtained. In Ekoporten some
of the nitrogen was probably lost as ammonia emissions during handling of the
samples (Paper III), while in Gebers approximately 32% of the phosphorus was
probably lost due to deposition in the system and the sampling equipment
(Andersson & Jensen, 2002) as noted earlier in studies by Vinnerås (1998).
In the double flush urine-diverting toilet system, some urine is impossible to
collect, as it is misdiverted. Flooding of the urine bowl and misuse of the toilet
mainly cause this. This has been identified as a major source of losses in several
studies of the urine-diverting system (Paper III; Jönsson et al., 1997; Jönsson,
Burström & Svensson, 1998; Vinnerås, 1998; Jönsson et al., 2000). After
compensation for misdiverted urine and nutrients lost in the sampling equipment,
the amounts of nitrogen and phosphorus collected corresponded well to the present
designing values.
Table 2. The collected amounts of urine concerning wet mass, dry mass, BOD7, COD,
nitrogen, phosphorus, potassium, copper, chromium, nickel, zinc, lead, cadmium and
mercury in the studies in Ekoporten and Gebers both measured (MD) and corrected for
misplaced urine (CD) per person and year compared to the Swedish designing values
(NV, 1995)
Parameter Unit Ekoporten Gebers Swedish Proposed new
MD CD designing values designing values
Wet mass kg 376 554 510 365 550
Dry mass kg 7 10 7.0 21 21
BOD7 g - - 1829
COD g - - 3723
N g 2500 3700 3830 4000 4000
P g 230 340 250 365 365
K g 800 1190 820 910 1000
Cu mg 1490 17.2 37 37
Cr mg 10.3 0.16 3.7 3.7
Ni mg 32.7 3,65 2.6 2.6
Zn mg 150 102 16.4 16.4
Pb mg 15.3 4.2 0.73 0.73
Cd mg 0.48 0.08 0.37 0.25
Hg mg 0.30 0.16 1.1 0.30
The amount of potassium in the Ekoporten urine mixture was considerably higher
than the expected amount according to the designing value (Table 2). In the study
in Ekoporten the potassium content in the urine was 30% above the designing
value and the content in the faecal water was almost 50% above the designing
value, but in the study in Gebers the collected amounts were somewhat lower than
the present designing value, both in urine and faeces. Several other studies of the
composition of human urine performed during the late 1990s also showed high
amounts of potassium compared to the other nutrients analysed and compared with
the Swedish designing values (Jönsson et al., 1997, 1998, 1999, 2000; Vinnerås,
1998, 2001). The main reason for the higher amounts of potassium is probably a
larger consumption of potassium, e.g. increased use of potassium-supplemented
salt. Therefore a proposal is made for an increase in the designing value for
potassium in urine to 1000 g p-1 y-1 (Table 2) This change also changes the ratio
between the potassium in the urine and the faeces to a level more reflecting the
expected physiological distribution of >70% in the urine and <30% in the faeces
(Berger, 1960).
The essential heavy metals; Cu, Cr, Ni and Zn, are present in the diet and are also
available in mineral supplements and are therefore influenced by the diet. These
metals also have a high presence in materials used in society. In Ekoporten, the
amounts of the essential heavy metals collected were higher than expected, while
the amounts in Gebers were about as expected. The high amounts in Ekoporten
were probably due to misplaced domestic cleaning water and corrosion of pipe
material (Paper I; III).
The present designing value is based on the average Swedish diet (NV, 1995) but
the amount of the investigated non-essential heavy metals (Pb, Cd and Hg) in
Swedish society has decreased during recent years (Lohm et al., 1997; Balmér,
2001) and this has probably also affected the dietary intake of these heavy metals.
However, the detected level of lead in the urine in Ekoporten was considerably
higher than the designing value but the levels of lead in all of the analysed
fractions in Ekoporten were high. The probable reason for the high levels was
deposition of lead from propeller plane fuel used at the local airfield.
The amount of cadmium collected in the urine in Gebers was only one fifth of the
Swedish designing value, while the amount collected in Ekoporten was
comparable to the designing value. For mercury, both of the analyses gave levels
considerably lower than the designing value, confirming the decreased flow of
mercury in the society. Due to the significantly lower amounts of cadmium
detected in Gebers and the low amounts of mercury detected in both analyses and
the fact that the given designing value for mercury was a detection limit, it was
proposed that the designing value for these two metals be lowered. The proposed
new level for cadmium was 0.25 mg p-1 y
-1 and for mercury 0.30 mg p-1 y
(Paper I).
The dry matter of faeces and toilet paper collected in Ekoporten, 12.5 kg p-1 y
was considerably lower than the designing value, 12.7 kg p-1 y-1 for faeces and 8.5
kg p-1 y-1 for toilet paper (Table 3). In Gebers a somewhat higher amount of faecal
dry matter, compared to Ekoporten, was collected, 10.0 kg, together with 8.5 kg of
toilet paper used. This indicates a lower excretion of faecal dry mass than the
designing value, which therefore is proposed to be somewhat decreased to 11.0 kg
per person and year (Table 3; Paper I).
The amount of toilet paper used in Sweden (8.5 kg DM p-1 y-1) (Anonymous, 1994)
corresponds well to the amount used in Gebers, which can therefore be seen as a
confirmation of the Swedish consumption (Paper I).
The wet mass of faecal matter at Gebers (72 kg p-1y-1) was twice the present
designing value, which corresponds to a water content of 88% at Gebers and 63%
in the designing value. Faeces collected from 20 adults had a water content of 77%
(Lentner et al., 1981). The present designing value for wet mass is probably
underestimated, so an adjustment is proposed to 51 kg p-1y-1 (Table 3).
The nutrients collected in Ekoporten and Gebers differ somewhat compared to the
present designing value. However apart from potassium, the values were not
significantly different from the expected amounts. The differences are probably
due to differences in the diet compared to the average Swedish diet.
Table 3. The collected amounts of faeces concerning wet mass (WM), dry mass (DM),
nitrogen, phosphorus, potassium, copper, chromium, nickel, zinc, lead, cadmium and
mercury in the studies in Ekoporten (En) both measured (MD) and corrected for
misdiverted material (CD); and Gebers where the amount of toilet paper used (TP) is
reported separately per person and year compared to the Swedish designing values for
faeces (NV, 1995) flushwater and toilet paper (Anonymous, 1994)
Unit Ekoporten Gebers Swedish Flush Toilet Proposed
CD CD TP Value water paper new value
WM kg 18700a 72b 8.9 36.5 18300 8.9 51
DM kg 12.6 10.0b 8.5 12.8 8.5 11
BOD7 g - 1223c
COD g - 1668c
N g 630 710 550 550
P g 126 250 183 183
K g 540 280 365 365
Cu mg 1060 628 400 400
Cr mg 68 47 7.3 7.3
Ni mg 110 81 27 27
Zn mg 4860 16900 3900 3900
Pb mg 460 13 7.3 7.3
Cd mg 6.4 5.7 3.7 3.7
Hg mg 2.8 3.2
23 3.3
aflushwater included
b faeces only, toilet paper (TP) excluded
c including TP
The amounts of essential heavy metals collected in Ekoporten were somewhat
above the expected amount but this could not be excluded as contamination from
external sources such as domestic cleaning water poured into the toilet. In Gebers,
the essential heavy metals in the collected material were fairly similar to the
designing value except for the amounts of zinc, which were probably due to
contamination from the zinc-galvanised pipes leading down to the collection bins.
Therefore no changes are proposed in the designing values for the essential heavy
In both of the investigated areas, the amounts of lead were significantly above the
values in the designing value. The reason for these high levels is not known but the
levels are too high to be related to the faeces itself. The high amounts are probably
caused by some kind of contamination. However, the levels of heavy metals
should be monitored more thoroughly to find their source and to see if the
consumption is as high as indicated by the measurements (Paper I).
The levels of mercury in both areas were significantly below the present designing
value, which was based on measurements from 10 people (Skare & Engqvist,
1992). A change in the designing value for mercury in the faeces to 3.3 mg p-1 y-1
is proposed (Table 3).
The flow of greywater at the three investigated sites was 38 m3 p
-1 y
-1 in
Ekoporten, 40 m3 p-1 y-1 in Gebers and 24 m3 p-1 y-1 in Vibyåsen (Table 4). These
flows were all considerably lower than the present Swedish designing value of 55
m3 p
-1 y-1 (NV, 1995). In a study performed by the Swedish EPA (NV, 1995),
greywater from an ecological village was investigated and the magnitude of the
greywater flow was in the same range as in our studies. One reason for the
diminishing water consumption is the use of water-conserving equipment. The
proposed new designing value for the greywater flow was 36.5 m3 p-1 y-1 (Table 4).
Potassium in the greywater at the three sites was also increased compared to the
Swedish designing value (Table 4). The reason for this was probably an increased
use of potassium, for example in liquid detergents and as a mineral supplement in
table salt. An increase in the designing value for potassium to 365 g per person and
year was proposed.
The major proportion of the heavy metal load to household wastewater was found
in the greywater fraction. Yet all the new data on heavy metals in greywater were
significantly lower than the present Swedish designing value, except for the copper
(Paper I). The present Swedish designing values for greywater were mainly based
upon measurements over seven days at sixteen apartments at the eco village
Tuggelite in Sweden (NV, 1995), corresponding to approximately 400 person
days. Data presented on greywater characteristics in this thesis correspond to
approximately 3050 person days, plus calculations on the sludge from 605000
people connected to the Ryaverket wastewater treatment plant in Gothenburg,
Sweden, over a 10-year period. Since discrepancies were revealed between the
present designing values and the new greywater data for all of the metals studied,
changes in the designing values for metals were proposed according to the
following (Table 4).
Copper is the only metal for which the designing value is proposed to be increased.
The concentration of Cu in the greywater is very sensitive to local conditions such
as corrosion of the water distribution pipes and service pipes. Data from our
studies in Paper I show chromium, nickel and zinc levels significantly below
present designing values. The new proposed designing values for Cr, Ni and Zn
were 365, 450 and 3650 mg p-1 y-1, respectively (Table 4).
The levels of lead were in all cases except for Ekoporten considerably lower than
the present designing value. A probable reason for the high levels of Pb in the
wastewater at Ekoporten was the nearby propeller plane airfield, where leaded fuel
still is used. The new proposed designing value for lead was 365 mg p-1 y-1. The
levels of cadmium (Cd) and mercury (Hg) in all the new greywater data were
significantly below the present designing value. For Cd the proposed new
designing value was 15 mg p-1 y-1 and for Hg 1.5 mg p-1 y-1.
Table 4. The collected amounts of greywater concerning wet mass (WM), dry mass (DM), nitrogen, phosphorus, potassium, copper, chromium, nickel, zinc,
lead, cadmium and mercury in the studies in Ekoporten both given as measured value (MD) and any misplaced domestic cleaning water (CD), Gebers,
Vibyåsen and Gothenburg per person and year compared to the Swedish designing values (NV, 1995)
Ekoporten Gebers Vibyåsen Göteborg Swedish Proposed new
MD CD designing value designing value
WM Kg 38000 40150 24236 - 55000 36500
DM Kg 21.6 14.6 15.1 - 29.2 20
BOD7 g - 7700 10100 - 10220 9500
COD g - 17500 14250 - 26280 19000
N g 613 510 230 - 365 500
P g 162 220 180 - 110 190
K g 1450 350 200 - 180 365
Cu mg 3010 3650 2365 1497 2740 <2190 2900
Cr mg 360 430 149 91.3 400 <1830 365
Ni mg 300 430 87 266 590 <1100 450
Zn mg 4800 6500 2252 1570 3290 <18250 3650
Pb mg 515 990 87 62.1 390 <1095 365
Cd mg 11.3 14.4 5.1 2.19 17.2 <219 15
Hg mg 3.8 3.8 1.1 0.37 <0.1 <21.9 1.5
Proposed new designing values for the Swedish wastewater
Based upon the measurements in Ekoporten, Gebers and Vibyåsen, combined with
the calculations on the sewage sludge in the Ryaverket wastewater treatment plant
in Gothenburg, it was possible to scrutinise the present designing value for the
composition of the wastewater (Paper I). The designing value for the solid
biodegradable waste is also included. As the last of the organic fractions from
households worth recycling, this fraction is also constantly under discussion for
introduction as a fraction in the wastewater by the use of waste disposal units in
the kitchen (Kärrman et al., 2001). The proposed changes to the designing value
are printed in bold (Table 5), the old designing values for the different fractions
can be found in Tables 2-4.
Table 5, The proposed designing value for composition of the different fractions of
household wastewater and biodegradable solid waste per person and year, proposed
changes are given in bold (Paper I)
Urine Faeces Toilet
Greywater Biodegradable
WM Kg 550 51.5 8.9 36500 80.3
DM Kg 21 11 8.5 20 27.5
BOD7 g - - - 9500 -
COD g - - - 19000 -
N g 4000 550 500 550
P g 365 183 190 104
K g
1000 365
365 82
Cu mg 37 400 2900 549
Cr mg 3.7 7.3 365 137
Ni mg 2.6 27 450 82.3
Zn mg 16.4 3900 3650 700
Pb mg 0.73 7.3 350 275
Cd mg 0.25 3.7 15 2.7
Hg mg 0.30 3.3
1.5 0.25
Faecal Separation (Paper II, III)
As detailed above, the major proportion of the nutrients is found in the urine,
which is a clean fraction with regard to heavy metal contamination. The next
fraction that contains a lot of relatively clean nutrients is the faeces (Figure 2; 3;
4). Compared to the urine, which has water-soluble nutrients, the faeces contain
both water-soluble nutrients and nutrients that are combined in larger particles not
soluble in water. Still, about 50% of the nitrogen and the majority of the potassium
in faeces are soluble in water (Berger, 1960; Trémolières et al., 1961; Guyton,
1992; Fraústo da Silva & Williams, 1997). Phosphorus is mainly found as calcium
phosphate particles, insoluble in water (Fraústo da Silva & Williams, 1997).
By using urine-diverting toilets, the majority of household nutrients can be
collected into the urine fraction (Jönsson et al., 1997, 1999, 2000; Höglund et al.,
1999). With no urine mixed with the flushwater, only small amounts of nutrients
will be found dissolved in the water. Therefore it is possible to use water to
transport the faeces to a unit that separates the water and a small volume, relatively
dry, faecal fraction. This can be collected before secondary treatment and reuse as
a fertiliser, while still making it possible to have the advantage of a flush toilet.
To determine possible length of the faecal transport before successful separation
and whether it was possible to have a slow dewatering of the faeces, the
biochemical activity was monitored by measuring the reduction and oxidation
potential in a faeces:water mixture at different availability of oxygen (Paper II).
The distribution of the nutrients between the solid and the liquid phase was also
analysed to see how the nutrients were lost to the liquid fraction upon separation,
over a filter with a pore size of 70µm.
When the faeces were immersed into the water, the redox potential was constantly
reduced even when air was bubbled through the solution (Paper II). The largest
drop in potential occurred when no oxygen was present. The speed of the change
in redox potential of the mixture implies high enzymatic activity degrading the
organic matter of the faeces, as observed also in the separated sludge in septic
tanks (Philip et al., 1993). The high biochemical activity was also noticed by the
increasing total concentration of ammonia, indicating a degradation of amino-
containing organic matter (Figure 8).
Time (min)
Ammonia/ammonium (mmol)
Figure 8. The distribution and the sum of ammonia over time in the two separated fractions
separated water (SepW) and separated faeces (SepS) given as an average for all three
Even if no significant difference was found between the amounts of nutrients lost
to the separated water fraction, there was a small tendency for higher losses at
higher levels of available oxygen and thus a higher redox potential. This was
expected due to the much higher levels of energy released when using oxygen as
an electron acceptor compared to other compounds used under anoxic and
anaerobic conditions. The high energy content in the chemical reactions when
oxygen is present gives more rapid reactions and thereby increased losses of the
elements studied due to degradation of the material.
The increased amount of ammonia found in SepW indicated an increased loss of
nutrients as a function of the time before separation. A lot of the nitrogen in faeces
occurs as organic nitrogen, which forms ammonia upon degradation (Trémolières
et al., 1961).
However, it was potassium that had the highest extraction rate of the elements
analysed (Figure 9). This is due to the form in which potassium occurs in the
faeces, i.e. mainly as free potassium ions (Berger, 1960; Guyton, 1992).
0 5 10 15 20
time (min)
Amount in SepS (%)
Figure 9. The average and the standard deviation of N, NH4, P and K remaining in separated
solids (SepS) after immersing the faeces in water for different periods of time.
Some increase in the loss of phosphorus was also observed but not as much as for
the two other elements. In faeces, phosphorus is mainly found as granular calcium
phosphate that seems to be captured in the 70µm filter used for the separation.
Some smaller particles and some organically bound phosphorus released by the
degradation of organic matter cause losses of nutrients in SepW.
This high biochemical activity results in losses both of elements released by
degradation of organic matter and by extraction when the contact surfaces and
contact time increase and a lot of the initially captured nutrients are eventually
dissolved into the water as in septic tanks (Philip et al., 1993). Thus, if the aim of
the separation is to recover faecal plant nutrients, it is important that the faeces and
the water are separated as soon as possible, since nutrient recovery decreases
rapidly with time.
Laboratory-scale evaluation of different separation techniques
(Paper II)
Four different separation techniques were evaluated during these tests, namely
flotation; sedimentation; filtration; and whirlpool, surface tension separation
(Paper II). Separation by flotation and sedimentation was only tested in small
initial tests. From those tests, it was apparent that for these methods the delay until
two separate fractions were achievable was too long, when compared to the tests
where the faeces were merged into water. These showed increasing rates of
nutrient losses with longer contact time between the faeces and the water.
Therefore, treatment by sedimentation and flotation were not considered as
possible alternatives (Paper II).
The separation with filtration and whirlpool, surface tension separation in an
Aquatron separator (Figure 10) was performed after a 1-metre vertical transport
and a 1-metre horizontal transport, connected by a sharp 90° bend. Faeces were
first applied to the pipe with a small amount of water and then the rest of the water
was applied to simulate an ordinary flush (Paper II). This procedure resulted in a
very turbulent transport in the pipe that broke down some of the structure of the
faecal particles and gave them a larger contact surface, thus increasing the losses,
from the separated faeces, of particles and nutrients during the separation.
Figure 10. The whirlpool, surface tension separator Aquatron and a technical description of
the function, FW=incoming Faecal Water.
The two separation techniques fully investigated gave similar separation of the
analysed elements, approximately 70%, with some differences in the distribution
between the elements (Figure 11). The difference in distribution of the elements
was probably due to differences in their forms and an effect of the separation
DM Ash Org N P K
DM Ash Org N P K
a b
Figure 11. The distribution of DM, Ash, Organic matter, N, P, K between the liquid and the
solid fraction when using (a) whirlpool, surface tension separation and (b) filtration.
The whirlpool, surface tension separation process separates the particles that are
unable to follow the walls of the separator, while particles that are suspended in
the water follow the separator walls to the separated water fraction. The material
lost mainly consists of suspended particles and, as can be seen in Figure 11a, the
percentage losses of all parameters were similar.
In the filtration, more of the potassium and less of the nitrogen was lost (Figure
11b). The probable reason was the increased contact time between the particles
and the water, compared to the whirlpool, surface tension separation. No visible
larger particles were detected in the separated water after the filtration (Paper III)
and the losses of dry matter and organics were negligible (Figure 11b). The lost
nutrients can therefore be assumed to be nutrients extracted from the particles
during transport and filtration.
During filtration, the filter cake functions as a supplement to the filter, but the first
material captured is washed by water passing through the filter. This effect,
together with the fast degradation and high activity in the faeces that were merged
into water, indicates the need for a rapid removal of the filter cake to avoid losses
of nutrients due to degradation and extraction.
Controlled pilot-scale separation of faeces and toilet paper by
using whirlpool, surface tension separation
The promising results from the laboratory-scale studies of the faecal separation,
using whirlpool, surface tension separation indicated that this method could be a
promising alternative for collecting the faecal nutrients in a small and dry fraction
and still using a flush urine-diverting toilet (Paper II). Therefore, a pilot-scale
system was built to investigate the potential of whirlpool, surface tension
separation with a toilet and 5-metre vertical drop, corresponding to a two-story
building, before the separation (Figure 12).
The system consisted of a toilet placed above a 5-metre vertical pipe connected to
two 45° bends, giving a gradual 90° bend. This was connected to a 1-metre
horizontal pipe with a 4% gradient linked to the whirlpool, surface tension
separator (Figure 12). The gradual 90° bend was chosen as a result of observations
in the laboratory-scale study (Paper II), where the sharp bend resulted in
disintegration of the faecal particles.
Figure 12. The pilot-scale whirlpool, surface tension separation system used for
investigating the effects on separation of water volume and toilet paper use.
Four different combinations of separated material and flushwater volume were
tested in the system:
1. Faeces flushed with four litres of flushwater.
2. Toilet paper flushed with four litres of flushwater.
3. Faeces and toilet paper flushed with four litres of flushwater.
4. Faeces and toilet paper flushed with six litres of flushwater.
On average, 50±5g of faeces and 27±3g of toilet paper were applied to the toilet
before flushing. In cases where both faeces and toilet paper were flushed, the
faeces were added first, followed by the toilet paper. Ten seconds after addition of
the material to the bowl, the toilet was flushed. Four series of five flushes were
performed for each material combination, giving a total of 20 flushes per
separation strategy. The separated mass was noted and weigh-proportional samples
were taken from every series. The samples were analysed for their dry matter
(105°C for >24h), ash (>550°C for >4h), nitrogen (Kjeldahl), phosphorus (ICP-
AES), potassium (ICP-AES) and sulphur (ICP-AES) content.
The amounts of flushwater used in separation strategy number 3 varied a lot
(±10%) (Table 6) due to the use of an extra flush to remove a blockage caused by
toilet paper in the pipe. Some tests were also performed where only 2 litres of
flushwater were used to flush 3×9g of toilet paper. In these tests, only every
second flush managed to flush the paper all the way to the separator, in the other
cases the paper became stuck in the initial bend in the toilet and needed another
flush to be transported to the separator.
Table 6. Amounts of material separated and collected matter for the different material
combinations per series of 5 flushes
Mtrl Mtrl in H2O in Mass SepS Mass SepW DM SepS DM SepW
Comb. kg L kg Kg kg kg
1 0.27±2.4% 19.1±2.0% 0.60±12% 19.0±2.4% 0.055±17% 0.008±93%
2 0.044±4.2% 20±1.6% 1.7±5.8% 18±2.7% 0.046±5.1% 0.0022±41%
3 0.29±3.5% 20±10% 2.0±7.5% 19±12% 0.10±21% 0.006±21%
4 0.31±3.2% 29±0.9% 2.7±12% 27±2.0% 0.11±5.4% 0.009±28%
The flushwater used was also analysed for its content of nitrogen, potassium and
sulphur. Therefore, the distribution of the nutrients between the two fractions is
given both according to the analysis results (m) and calculated results (c) in which
the nutrients in the water used for flushing were deducted (Table 7; 8).
The concentrations of the elements collected in the separated water differed
greatly, giving a high standard deviation. However, in the separated faeces the
collected amounts were more stable, both within each material combination and
between the different material combinations (Table 7). The concentration of the
elements nitrogen and potassium in the water used for flushing corresponded to
approximately 10% of the total amounts, while for sulphur approximately 30%
came from the water used. These initial concentrations are important to bear in
mind when evaluating separation systems by analysing the two separated fractions.
There were two main effects of including toilet paper in the separation in addition
to faeces, namely an increase in the total mass and the amount of dry matter
collected in SepS (Table 6; 8). The effect of the collected mass was that when no
paper was used, the amount of material collected in SepS was only 3% of the total
mass. When toilet paper was included, the mass was between 8.5% and 9.6% due
to absorption of water by the paper. Thus, the effect of collected volume was
highly influenced by the toilet paper, mainly due to its water absorption capacity.
Therefore, the number of flushes with toilet paper should be kept to a minimum.
Paper used on occasions other than defecation should either be collected in a bin
beside the toilet or in the rear bowl of the toilet and only flushed away when
The amounts of collected and analysed elements in the SepS fraction after
correction for the content in the flushwater were higher when smaller amounts of
flushwater were used (Table 8). The probable reason for this is an increased
disintegration of the faecal particles due to the larger water volume, but it still
seems to be possible to collect approximately 85% of the faecal nutrients in this
kind of system.
Table 7. Amounts of nitrogen, phosphorus, potassium and sulphur in the two fractions of separated water (SepW) and separated faeces (SepS), per series of 5
flushes, with both the measured value (m) and the calculated value (c), in which the initial content of the element in the flushwater was deducted
Material N SepS N SepW P SepS P SepW K SepS K SepW S SepS S SepW
comb. mg mg mg mg mg mg mg Mg
2m 3500±16% 760±46% 1200±11% 215±44% 1000±13% 520±16% 330±8.9% 290±10%
2c 3500±17% 310±110% - -
1000±13% 330±25% 320±9.0% 76±36%
3m 3300±22% 710±38% 1100±19% 180±18% 900±22% 470±11% 320±23% 260±8.5%
3c 3300±22% 270±81% - -
880±22% 280±13% 300±24% 69±3.2%
4m 3200±11% 1350±44% 1300±17% 240±9.4% 880±5.5% 610±7.3% 300±12% 380±1.8%
4c 3100±11% 710±83% - -
850±5.9% 340±15% 270±14% 88±13%
Table 8. Fractions (%) of the flushed matter and elements recovered in the fraction of separated faeces depending on separation strategy used. For the
elements nitrogen, potassium and sulphur, both the value measured (m) and a calculated value (c), in which the initial influence of the flushwater was deducted
are shown
Mtrl. comb Mass DM N (m) N (c) P K (m) K (c) S (m) S (c)
1 8.5% 95% - - - - - - -
2 3.0% 87% 82% 92% 85% 66% 75% 53% 81%
3 9.6% 94% 82% 92% 86% 66% 76% 55% 81%
4 9.1% 92% 70% 81% 84% 59% 72% 45% 76%
The difference in separation between the elements analysed, and how the
separation was influenced by the flushwater volume used, is probably due to the
different forms of the elements in the faeces, e.g. nitrogen is mainly bound in
organic matter (Trémolières et al., 1961), phosphorus is mainly bound as granular
Ca(PO4) and potassium occurs mainly as free ions (Fraústo da Silva & Williams,
1997). The tendencies for N, P and K were the same as in the laboratory-scale
studies in Paper II. However, more nutrients were collected in SepS in these
studies (85%) compared to the laboratory-scale studies (70%) (Paper II). The main
reasons for the difference are probably that in the laboratory-studies no correction
was made for content of the elements in the flushwater and higher disintegration
occurred due to the sharp 90° bend in that study. It seems more important to have a
gradual bend than to keep the height of the vertical drop very small.
If the system is built correctly, it appears possible to use whirlpool, surface tension
separation in buildings with several floors keeping the separation equipment on the
ground floor or in the basement. Larger systems mean longer transport and higher
disintegration of the particles, which leads to increased loss of nutrients. However,
it seems possible to collect the main proportion of the nutrients in the drier fraction
of separated faeces also in correctly built fairly large systems.
Full-scale separation of faeces in the Ekoporten block of flats
(Paper III)
In the Ekoporten block of flats, all 18 apartments are equipped with urine-diverting
toilets and the faecal bowl of the toilets is connected to two parallel whirlpool,
surface tension separators (Figure 10) in the basement. The separated solids are led
to a continuous composting reactor, where they are treated together with the
biodegradable household waste using pelletised sawdust as an amendment
(Paper III).
The initial analysis of the collected material showed poor recovery of the nutrients
from the faecal fraction. However, after correction for misdiverted urine according
to the weight of the fraction, the separation of the faeces appeared fairly good. In
the fraction of separated solids (SepS), 13% of the total mass, 37% of the dry
matter, 59-73% of the nitrogen, 58% of the phosphorus and 45% of the potassium
were recovered (Figure 13). The range in nitrogen recovery depended on the
correction for lost nitrogen from the urine mixture during handling of the samples
before analysis.
Comparing the laboratory-, pilot- and full-scale studies reveals that much more
nutrients were captured in the two small-scale studies than in the full-scale one. On
average 55-60% of the nutrients were collected in the full-scale study (Figure 13;
Paper III), 70% in the laboratory-scale separation studies (Figure 11; Paper II) and
80-85% in the pilot-scale separation studies (Table 8).
Mass DM N P K
Figure 13. The distribution of mass, dry matter, nitrogen, phosphorus and potassium
between the two separated fractions separated solids (SepS) and separated water (SepW) in
the Ekoporten block of flats after correction for the misdiverted urine mixture and the
nitrogen loss during handling.
The reason for these large differences between the different studies can be related
to several circumstances. In the collection in Ekoporten (Paper III), a lot of urine
was collected (32%) in the faecal fraction and the correction for this gave
uncertain values, especially for nitrogen. At first, the difference was linked to the
size of the system and the large drop in Ekoporten (4 stories) compared to the 1-
metre drop in the laboratory (Paper II). However after the pilot-scale study
(above), where the separation was performed after a 5-metre vertical drop where
the collection of the nutrients into SepS was higher compared to the laboratory-
scale experiments, it seemed as though other factors had a larger impact on the
efficiency of separation than the vertical drop.
According to the producer of the whirlpool, surface tension separation system
Aquatron, the system in Ekoporten is not optimal for the separation due to the
incorrect gradient of the pipes connected to the separator (Åkesson, 2002).
However, one factor that seems to have a large influence is the bend connecting
the vertical and the horizontal pipes. In the laboratory and the full-scale studies,
the bend was a sharp 90° angle but in the pilot-scale study two 45° bends were
used to give a more gradual bend, which decreased the impact and reduced the
disintegration of particles and the loss of nutrients to the separated water.
Important issues when using faecal separation
The pilot-scale tests (above) showed during the adjustments of the flushwater
volume that it was possible to reduce the amount of water collected in SepS to as
little as 0.2% when only water was flushed. The investigation in Ekoporten
showed that the majority of the heavy metals from domestic cleaning water poured
into the toilet ended up in SepW. However, it is still important not to pour the
domestic cleaning water into the toilet in this type of system (Paper III). The high
concentration of heavy metals in the cleaning water and the high density of metal
particles, which increases their separation, combined with the comparatively low
heavy metal content in the faeces indicates that domestic cleaning water should be
disposed of somewhere else if the solids collected in the toilet are to be recycled.
The whirlpool, surface tension faecal separation system is fairly insensitive
towards domestic cleaning water, as the majority is found in SepW (Paper III).
Other systems where all material from the toilet is collected unseparated, i.e. both
vacuum and low flush blackwater systems are much more affected by misplaced
contaminated water.
The effect on volume collected in SepS of the number of flushes depended on
whether any other material also was flushed. If toilet paper was flushed, about 8-
10% of the water became separated into the separated solids fraction (SepS), but if
only water was flushed, the effect was not noticeable. Therefore, toilet paper used
on occasions other than defecation should if possible be disposed elsewhere, either
collected in the rear bowl and flushed at need or collected in a bin beside the toilet.
The effect of this is considerably less water to handle in the SepS fraction.
The effect of the flushwater volume used is relatively small (see above). When the
volume of the flushwater used increased by 50%, the increase in SepS was just a
few percent (Table 8). Of greater importance here is the risk for blockages when
insufficient volumes of flushwater are used. However, somewhat larger nutrient
losses occur when the flushwater volume is increased, probably due to higher
disintegration of the faecal particles and thereby higher extraction rate and a larger
amount of small non-separable particles.
For optimal function of faecal separation as regards nutrient capture, the urine has
to be diverted. The faecal separation depends on larger particles and the nutrient
content in the urine is soluble in the water and is therefore not possible to separate
to SepS. By using a urine-diverting toilet, the major proportion of the nutrients
from the urine can be captured in a separate liquid low volume fraction (Jönsson et
al., 1997, 1999, 2000; Vinnerås, 2001).
One alternative to faecal separation that guarantees that all nutrients are collected
separately without any water addition is to use dry urine-diverting toilets (Figure
6). In the Western world of today, where people are used to water closets, it would
be hard to reintroduce dry toilets. In large parts of the world, however, any kind of
toilet would be an improvement, compared to open field defecation. Using urine-
diversion combined with faecal separation makes it possible to recover the major
proportion of the nutrients from households in unpolluted fractions and still use the
advantage of a water closet.
If filtration is used instead of whirlpool, surface tension separation, the filters
should have a short detention time. Filters with long detention times lead to large
losses of nutrients due to degradation of the material combined with the washing
of the filter cake by the following separations (Paper III).
Disinfection of faeces (Papers IV, V)
To ensure safe recycling of plant nutrients from sewage products, potential
pathogens have to be inactivated. Otherwise there is a large risk for transmission
of diseases when recycling wastewater products.
Disinfection of human urine is possible by storage. This is mainly due to a
combination of low content of organic matter, high amount of free ammonia and
high pH. After six months of storage at 20°C, the urine can be used as a fertiliser
even for foods consumed raw by humans (Jönsson et al., 2000; Höglund, 2001).
For other wastewater products, such as faeces and sewage sludge, storage is not an
efficient disinfection alternative. The differences in these materials compared to
the urine are lower pH and less ammonia. Thus, the amount of free ammonia is not
high enough to have satisfactory antibiotic effect towards the microorganisms
(Werkerle & Albrecht, 1983; Allievi et al., 1994; Gantzer et al., 2001). The large
content of organic material also affects the inactivation of pathogens negatively by
providing both physical protection and as energy for the organism metabolism.
When there is a high water content of the materials anaerobic degradation occurs.
This degradation, without oxygen as oxidising agent, will result in rest products
that have negative effects to the environment (Flodman, 2002).
Two high-tech alternatives for disinfection of sewage products such as faeces and
sludge with dry matter contents between 1% and 10% dry matter are liquid
composting and thermophilic anaerobic digestion. These systems have to be used
on a large scale to be economically feasible alternatives, and in a lot of cases when
the systems are used for toilet waste, additional organic matter, such as solid waste
and manure, has to be added. The energy requirement of these systems is also
high, e.g. for wet composting, the aeration is the single most energy-demanding
part of the sewage system (Kärrman et al., 1999; Skjelhaugen, 1999a-b).
Alternative systems to those available today have to be developed for disinfection
of biodegradable waste products from society, to make it possible to recycle them
in a safe way. The alternatives investigated in this thesis were dry thermal
composting and chemical disinfection using ammonia and a peracetic acid/
hydrogen peroxide mixture. One main advantage with these methods is their scale
independence, which makes them viable alternatives both for treatment of waste
products from single households and treatment of the sewage sludge from large
treatment plants.
Thermal composting of separated faecal matter (Paper IV)
Investigations into the disinfection of faeces by composting have been performed
both in Ethiopia (Karlsson & Larsson, 2000) and in Mexico (Björklund, 2002b).
These did not show any significant rise in temperature apart from an
approximately 10°C increase above the surrounding temperature during the first
days of treatment. Different reasons for this small rise in the temperature were
identified, such as insufficient insulation and low energy content due to addition of
ash during collection of the faeces.
To see if it was possible to get high temperatures (>50°C) by dry composting of
faeces, a laboratory-scale study and a pilot-scale study was carried out with
different mixes of faeces, urine and food waste combined with amendment
(Paper IV). The laboratory studies were performed in 1-litre Dewar vessels (Figure
14), and the temperature and the degradation of organic matter were monitored
(Paper IV).
Figure 14. One of the Dewar vessels used for laboratory-scale experiments on thermal
Four different mixtures were tested (Table 9) and amendment (chopped straw) was
added to all of the mixtures. The amount of amendment used strongly depended on
the water content in the rest of the material. Amendment (straw) was added to
reach an initial dry matter content of approximately 35%.
Table 9. The composition of the material composted in the different vessels during the
laboratory study (% of wet weight and % of dry matter) F=faecal matter, FF=faecal matter
+ food waste, FFU=faecal matter + food waste + urine & FU=faecal matter + urine
Vessel Amendment Faeces Artificial food waste Urine
ww dm ww dm ww dm ww dm
F 15% 41% 85% 59% - -
FF 10% 20% 65% 33% 25% 47% -
FFU 15% 38% 6% 4% 24% 58% 55% 1%
FU 28% 93% 7% 6% - 65% 1%
Pelleted dog food was used as artificial food waste. The advantage of using
pelleted dog food as artificial food waste in this kind of test is the assurance of
getting a homogeneous mixture in all of the test vessels. In these tests, the dog
food was left as pellets to compensate for the lack of cellular structure present in
regular food waste, which results in a lower direct availability of energy and
nutrients in the material. Other researchers have also used dog food as artificial
food waste (Nagasaki et al., 1996).
The composition of the dog food was similar to that of the standardized food waste
mix used by Eklind & Kirchmann (2000), which was a mix of ground potatoes
(65% of dry matter), carrots (15%), meat meal (13%) and bone meal (7%). The
major difference between the dog food used here and the waste used by Eklind &
Kirchmann (2000) was the ash content, which was 5.9% and 19% respectively. It
is 13-37% in regular biodegradable household waste (Sonesson & Jönsson, 1996;
Eklind et al., 1997). On the other hand, the ash content of the kitchen waste
collected at the site of the pilot-scale testing was lower (2%) than that of the dog
food used here (Paper IV).
0 1 2 3 4 5 6 7 8 9 101112131415161718
Time (days)
Temp (°C)
Figure 15. The temperatures during composting of the different mixtures in the laboratory-
scale study, amendment was added to all mixtures (Table 9), FF (Faeces, Food waste), FFU
(Faeces, Food waste, Urine), F (Faeces), FU (Faeces, Urine) and, RT (Room temperature).
In the vessel with faeces, urine and straw amendment (FU), the increase in
temperature was only 10°C during the first days (Figure 15), as in the tests in
Ethiopia (Karlsson & Larsson, 2000) and Mexico (Björklund, 2002b). The high
water content in the latrine (urine + faeces) resulted in high amounts of
amendment needed (28% of wet weight and 93% of the dry matter content) to
reach an initial dry matter content in the mixture of 35%.
According to Equation 2, the energy in the material should be sufficient to reach
elevated temperatures if we assume a biological degradation of faeces and urine
comparable to that of pig manure (58%), and of straw 55% (Haug, 1993). That
degradation is for a time span of 120 days and the assumption can be made that the
degradation during the 20 days in this investigation was smaller and thus smaller
amounts of heat were produced. The measured degradation of the organic matter
was 10% and when used in Equation 2, this gave a ratio of water content to
biological volatile substance of 23, which indicates that there was insufficient
energy present in the mixture to vaporise the water. The small increase in
temperature showed that this kind of mixture is not a suitable substrate for thermal
disinfection by composting.
The mixture with faeces and straw only (F) had a temperature development
comparable to the one with faeces, food waste and straw (FF). After 24 hours, the
activity in the faecal and straw mixture (F) more or less stopped and the
temperature decreased from over 40°C down to 25°C, probably due to low water
content. After watering and mixing, the temperature rapidly increased at day 4 up
to over 50°C (Figure 15), indicating high energy content, and the degradation of
the organic matter was 21%. However, the pure faecal matter was very sticky and
thus hard to keep structured and aerated.
The mixture of food waste, faeces, and urine (FFU) gave a constant high
temperature (Figure 15) and the second highest degradation (39%). One of the
major problems when including urine into the composting substrate is the
increased water content that has to be compensated for by addition of amendment,
resulting in very large volumes to be treated. The urine, with its high nitrogen
content, also increases ammonia emissions due to the high temperatures and the
high pH.
The mix that gave the highest temperature (>65°C, Figure 15) and the highest
degradation of organic matter (53%) was faeces and food waste (FF). The
temperature was over 50°C for more than eight days, indicating a high and stable
activity. However, the temperature fluctuated a lot during this time (Figure 15).
The main fluctuation was probably caused by the small mass that was treated
(approximately 400 g wet weight). Thus, it is important to mix the material often
and to reset the water content regularly.
The substrate mixture that gave the highest temperatures and degradations was the
faeces and food waste mixture (FF) and therefore this mixture was used in the
pilot-scale experiment. The only difference was that the amendment was changed
from straw to matured compost and the food waste used was regular food waste
collected on site (Paper IV).
The pilot-scale composting was performed in a 90-litre bin, insulated with 200 mm
of plastic foam on all sides (Björklund, 2002a). The aeration was secured via a
ventilation hole in the lower part of one side. The temperature in the compost
rapidly increased to approximately 40°C and then remained stable until day 11,
when the compost was mixed and the temperature increased to over 60°C (Figure
A temperature gradient from the middle of the compost towards the walls of the
bin was identified (Figure 16). This emphasises the importance of using sufficient
insulation so the influence of the outdoor temperature is kept as small as possible.
The insulation is important even when composting in high temperature climates if
thermal composting is desired. Even with the large amount of insulation used here,
200 mm of plastic foam; the material close to the ventilation hole probably did not
reach high temperatures.
0 5 10 15 20 25 30 35
Time (days)
Temp. (°C)
Figure 16. The average temperature every 10 minutes in the middle of the compost (Tm), at
one of the walls of the compost (Te) and in the surroundings (To) for the pilot-scale
experiments (Paper IV).
The importance of keeping a high temperature throughout the reactor is to ensure
that all of the material is thermally disinfected (see below). To avoid these lower
temperature zones, the ventilation air has to be preheated somehow. This can be
done in a heat exchanger for the outgoing air, which can simultaneously be treated
against emissions of odours and ammonia losses by condensing its moisture. In the
condensated liquid the majority of odour agents and ammonia will also be
captured (Haug, 1993; Skjelhaugen, 1999b).
Another way to preheat incoming air is by placing organic matter with a higher
hygiene status, e.g. food waste, closest to the air inlet and using its biological
activity to increase the temperature of the incoming air. Upon turning, the material
will have to be moved to another cell, which has been similarly prepared with
fresh organic matter closest to the ventilation hole, so no un-disinfected material
ends up in the low temperature zone and thus escapes thermal disinfection.
Large-scale composting reaches disinfection temperatures fairly easily due to its
temperature stability deriving from the large mass. The monitoring does not have
to be as intense as in the case of the small-scale composts. However, composting is
a biological process and the more closely it is monitored and controlled, the easier
it is to get a well functioning process.
The degradation of organic matter in the composts was high. Even though the
duration of the treatment was short, 19 days for the laboratory-scale experiments
and 35 days for the pilot-scale experiments, the degradation was 53% and 73%,
respectively. The general degradability of organic waste and faeces is between
55% and 80% (Haug, 1993), so one can assume that the stability of the material,
especially from the pilot-scale study, was high and that the material was well
suited for use as a soil conditioner. In addition, there is unlikely to be any
reheating of the material on application to soil.
In many cases, ash is added to the faeces during the collection in the toilet, a
practice especially common in developing countries. The ash dries out the material
and increases its pH. This prevents smells and flies and gives some inactivation of
pathogens, but not a complete disinfection. One other effect from this is that the
concentration of organic matter decreases, leaving less energy available to increase
the temperature. This is probably one of the reasons why high temperatures were
not reached in the tests on composting of faeces in Ethiopia (Karlsson & Larsson,
2000) and in Mexico (Björklund, 2002b). To achieve sufficiently high
temperatures, a high-energy amendment has to be added to the material, for
example fruit peelings. The compost also has to be well insulated. However, this
has to be investigated further to confirm mixtures and functions.
Disinfection of faeces
By using Equation 1 and the equations for the time required for inactivation for
different pathogens given in Table 1, it is possible to calculate the safety margins
for total die-off of different pathogens, i.e. where no more viable organisms should
be found in the laboratory-scale and the pilot-scale composts (Table 10). The
lowest safety margin was found to be the margin for the Enteroviruses, where the
faecal and food waste mix (FF) in the laboratory-scale experiment had a safety
margin of 130 times total inactivation, which can be considered an extremely high
margin. The mix of faeces, food waste and urine (FFU) did not reach the same
elevated temperatures as FF. The safety margins were therefore much smaller, for
Enteroviruses only 8 times and for salmonella and Ascaris only 14 and 15 times,
respectively. However, this is still a reasonable margin of safety when handling
this material.
In the pilot-scale study, the temperature was high during a longer period (Figure
15), even at the wall, compared to the lab-scale experiments (Figure 16). Also in
this case, the calculated safety margin for inactivation of pathogens was smallest
for Enteroviruses, 37 times. For the other pathogens, for which a limit of
inactivation was given by Feachem et al. (1983), the limit was fulfilled more than
100 times, e.g. for Salmonella, Shigella and Ascaris, the calculated safety margin
was larger than 100, for Taenia eggs and Schistosoma, the safety margin was more
than 1000 and for Enteric hystolica and Cholera the safety margin was larger than
This shows, as Figure 7 indicates, that the organisms least affected by the
treatment are Enteroviruses and Ascaris. Salmonella can also be considered to be
quite resistant towards thermal treatment. Therefore, when estimating the safety
margins as in Paper IV or when using organisms for thermal deactivation or
evaluating existing processes, these three organisms or organism categories, of the
above mentioned (Table 1), should if possible be included in the evaluation. Other
organisms can be included but as shown in Figure 7, the upper temperature/time
limit is set by Enteroviruses from approximately 51°C, and below that temperature
the limit is set by Ascaris, but the limit for no viable Salmonella is in the same
region during the temperature interval of 45-65°C. Of the other organisms
mentioned, the inactivation at these temperatures is so high (Table 10) that only if
the organisms are of special interest do they need to be studied more closely in
thermal composting.
Table 10. The safety margin for disinfection of the faecal compost as the number of times
the limit of no viable organisms found was fulfilled in the two most prominent mixes tested
in the lab-scale(FF=Faeces+Food waste; FFU=Faeces, Food waste & Urine) and the wall
and the middle of the pilot-scale reactor
FF FFU Wall Middle
Enteroviruses 130 8 37 110
Cholera 700 000 3 000 170 000 720 000
Salmonella 670 14 170 600
Shigella 1 800 50 480 1 600
Enterohystolica 650 000 540 150 000 800 000
Ancylostoma 2 300 70 610 2 100
Ascaris 2 300 15 550 2 200
Schistosoma 24 000 200 5 900 23 000
Taenia 4 100 120 1 100 3 600
This experiment showed that by mixing faeces, food waste and amendment, it is
possible to reach high temperatures, thus securing the disinfection of pathogens in
the material. Spores of bacteria will probably not be affected by these
temperatures, although the fluctuation in temperature arising from mixing and
adding water is reminiscent of a tyndalisation process that will affect the number
of surviving spores (Stanbury et al., 1995). However, there is still a risk for
regrowth of pathogenic bacteria if any survive the treatment or re-contaminate the
Unfortunately the temperature of the compost material directly at the air inlet was
not recorded. The temperature here was probably lower than in the rest of the
compost. By using Equation 4 it is possible to calculate the number of times the
material has to be mixed to attain a certain level of disinfection (Haug, 1993). The
equation assumes that you have total inactivation in the high temperature areas,
that the material is completely mixed in turning and that no inactivation or
regrowth occurs in the low temperature areas.
When low temperature zones are present, the pathogens will not be deactivated
within all material and there will therefore be an increased risk for regrowth of
pathogenic bacteria in the low temperature zones after thermal composting that has
to be considered. However, the risk for regrowth in composted material is smaller
compared to undegraded material (Sidhu et al., 2001) so composting will decrease
the risk even if some bacteria survive in the low temperature areas.
The high safety margins show that well functioning thermal composting is an easy
method for disinfection of material that can be used as a fertiliser and soil
conditioner with very small risks for transmission of diseases, provided that all the
material is disinfected and that the handling of the raw material is done with care.
Regardless of the efficiency of the composting process, one major risk for
transmission of disease is contact with the raw material when starting the compost
and with not fully disinfected material upon turning it. Management of this risk
requires good working practices, worker information and possibly some degree of
automation. Since one of the major causes of Ascaris transmission is contact with
raw faeces, this is an important factor to consider when planning this kind of
system. Precautionary measures have to be taken by informing the users of this
system and those who handle the matter, to avoid transmission of diseases.
The use of a closed compost reactor where the turnings are mechanised can
decrease the risk for transmission of pathogens. One extra advantage with a closed
reactor is the possibility to capture the energy of the outgoing air by heat exchange
and condensing the water. In the condensed liquid, up to 85% of the ammonia
emissions can be captured together with smelly organic matter (Beck-Friis et al.,
2001). Thus, both the captured heat energy can be used and the effects of ammonia
emissions and odour agents on the surroundings can be reduced.
After treatment, the material will be at risk for re-infection by pathogens. To
avoid this, the equipment used to move the disinfected material and the place
where the matter is stored should be free from pathogens. Otherwise, there will be
a risk for regrowth of pathogenic bacteria.
Chemical treatment of separated faecal matter (Paper V)
The characteristics of the two chemical disinfection methods tested were
completely different. The peracetic acid is a rapid disinfectant that has its main
effect within one hour after addition, while urea addition has a slower, and
delayed, effect since the added urea has to be degraded to ammonia before
disinfection starts.
Treatment with peracetic acid (PAA)
The main characteristic of PAA is its high oxidative activity. This high activity
results in reactions with all kinds of organic matter decreases. Thus the content of
organic matter can be a damper on the inactivation of organisms. Content of
organic matter can also result in production of foam. During these treatments, the
volume of the material was doubled due to foam production.
The effect of PAA on microorganisms is that surface proteins are denaturised,
mainly by destroyed sulphur- and sulhponic-bonds. As PAA is an oxidising agent,
upon reaction it is reduced to acetic acid (Equation 10) and the hydrogen peroxide
to plain water (Equation 11). As the chemicals are reduced, the sanitation ability is
consumed as can be seen in Figure 17, where there was no significant difference in
numbers of organisms for the Salmonella phage between one hour and five days.
For E. coli and Enterococcus spp the numbers of organisms increased by regrowth,
to a level higher than the initial concentration, as the active substances were
degraded. Also, the otherwise resistant clostridia spores were reduced by the
treatment (Figure 17).
Time (days)
Number of organisms (log)
E. coli
Figure 17. Reduction in viable organisms 1h and 5 days after treatment of faecal matter with
0.15% peracetic acid.
Due to the short period of active disinfection of the faecal matter, another test was
performed by increasing the concentration of PAA to 0.5%, 1.0% and 1.5% of
active substance in the final mixture. The only time of treatment monitored for
these concentrations was 0.5 days (12 hours). A significant increase in disinfection
of the monitored organisms was recorded as the concentration of the active
substance was increased. At 0.5% PAA it was still possible to determine viable
organisms, while at the higher concentrations no living organisms were detected at
Table 11. Reduction compared to the initial concentration in log10 of microorganisms on
addition of PAA to levels of 0.5%, 1.0% and 1.5% of active substance, > is when no viable
organisms were found and thereby the reduction is larger than the detection limit
Organism Initial conc. PAA0.5% PAA1.0% PAA1.5%
S. Typh. 28B 7.97 3.82 >5.97 >4.97
Clostridia 4.53 2.4 >3.53 >3.53
Enterococcus 6.06 3.87 >4.06 >4.06
E. coli Detected ~6-7 Detected Not detected Not detected
At these high concentrations of added PAA, it can be considered very safe to
recycle the treated faecal matter. The recommendation from the Swedish Board of
Agriculture for treatment of manure in farms with Epizootic diseases is that an
application of 4.5-6.4% of PAA (JV, 1997) seems to be the required dosage to
achieve a sufficient reduction in viable pathogenic organisms for safe recycling.
An additional effect from the PAA treatment is a significant smell reduction, even
at the lowest added concentration. It appears that PAA is an alternative for this
kind of material with such high organic matter content when the need for
disinfection is high, i.e. when there is a large risk for epidemics, e.g. in refugee
camps. However, it seems to be more efficient as a disinfectant for liquids and
should therefore preferably be used for material with a low organic matter content.
Treatment with Urea
Compared to PAA, urea treatment is a slow method of sanitation. The urea itself is
not toxic and is actually found in many body lotions and toothpastes, but when it is
degraded to ammonia and carbon dioxide by the enzyme urease it becomes
antibiotic. Due to the need for degradation of the urea, this treatment requires
about one hour of incubation before it is actually active as a disinfectant.
If the pH is buffered at a neutral value (~7) the sanitary effect from the ammonia
produced will be negligible as the main effect is from uncharged ammonia (NH3),
which only occurs in low concentrations at that pH.
When the pH increases, the ammonia produced has an effect. In the tests
performed in Paper V, where 30 g of ammonia nitrogen equivalents as urea were
added, the pH increased to above 9.2 within one hour. The disinfection effect
towards E. coli and Salmonella was rapid, and the D-value (decimal reduction
value) for inactivation of E. coli, when including the time for degradation of the
urea, was >0.6 days as no viable organisms were detected
(<1 log10) within 5 days of treatment (Figure 18; Paper V).
The Enterococcus spp were more resistant towards the ammonia with a D-value of
3 days and no viable organisms were detected within 21 days of treatment (<2
log10) (Figure 18). The last bacteria group analysed was spore-forming clostridia,
which did not show any significant reduction by the treatment.
0 102030405060
Time (days)
Reduction (log)
E. coli
Figure 18. Reduction in viable organisms 1h, 5 days, 20 days, and 50 days after treatment of
faecal matter with 3% of urea nitrogen.
The very ammonia-resistant bacteriophage Salmonella typhimurium 28B phage
was used as a model organism for viruses. The D-value for the phage was <7.5
days (Figure 18; Paper V). As this phage has been shown to have double the
resistance of Rotaviruses and 20 times the resistance of poliovirus, the main
proportion of the pathogenic viruses should be inactivated within two months of
Ascaris suum eggs were used as an indicator for parasites and a reduction in the
viability of the nematodes was detected. Within five days of treatment, the
viability had decreased by 15% compared to the controls. After 50 days of urea
treatment no viable A. suum were detected. The faecal matter itself also seams to
be enough for significant inactivation of the A. suum viability as only 0.2% of the
cysts were viable. These lethal conditions were reduced by the PAA treatment as
no significant difference in the viability was detected between day five and day 50
at 0.15% of added PAA.
The treatment of separated and collected faecal matter with urea seems to be an
efficient and safe way of achieving hygienically safe material within 2 months of
treatment at room temperature. At the same time, the fertiliser value of the faecal
matter will increase as the added urea nitrogen is still available as a plant nutrient
after the treatment, together with the original content of nutrients in the faecal
Risk of regrowth after disinfection treatments
One of the major risks after disinfection is the risk for regrowth of bacteria. Even
if no organisms are detected, it is possible, if any organism survives, that regrowth
of bacteria will occur as an effect of changes in the environment in the material.
Another potential risk is the post-treatment handling of the material before
application. If the treatment effect diminishes, there will be a risk for seeding of
pathogens by contact with untreated matter in the post treatment handling.
When thermal treatment such as pasteurisation is used, the stability of the material
is unaffected or decreased. Thus, there is a large risk for regrowth of pathogenic
bacteria if present due to the presence of available organic matter and the absence
of competing organisms. If the thermal treatment is performed by composting
instead, the risk for regrowth is smaller, but still present (Sidhu et al., 2001). The
risk for regrowth is reduced both by competing, non-pathogenic organisms, and a
reduction in readily available organic matter (Paper IV).
When PAA is used for disinfection of faecal matter, the treatment of the material
have to be performed within a few days before application to avoid problems with
regrowth of bacteria. In the treatment with 1500 ppm PAA, E. Coli had regrowth
to one log10 more than the initial concentration within five days after the
application of PAA (Paper V).
When using disinfection systems in which the disinfection agent is removed (e.g.
heat) or consumed during the treatment (e.g. PAA), it is therefore important that
the material be applied to arable land as soon as possible after the treatment.
Another approach to the disinfection is to use a disinfection agent that is not
consumed during the treatment. One example of this is the use of high pH, e.g. by
applying ammonia to the material.
The risk for regrowth after the ammonia treatment is small, as the effect of
ammonia does not decrease during the treatment provided that there is no change
in pH and temperature and no ammonia loss by air emissions. After urea
application, the enzyme urease degrades the urea to ammonia and carbon dioxide.
The ammonia produced increases the pH and under these conditions it has a high
antibiotic effect. If the treatment is performed in a closed container (no ammonia
loss), there is no risk for regrowth of bacteria (Paper V).
If pathogens are added to the material during the handling, the environment will be
as lethal as for pathogens present in the material from the start. In all cases of
disinfection treatment, there is a risk for surviving pathogens if the material is non-
homogeneous or if the treatment is non-homogeneous so that untreated volumes of
the material remain where organisms can survive. However, if the effective factor
remains in the material until application there is no risk of regrowth within the
treated part of the material and if a chemical treatment is used, the risk for
untreated parts occurring decreases with time as the chemicals strive for
equilibrium within the material.
Comparison of sewage treatment strategies
A comparison was performed for two parts of the system, the primary treatment of
toilet water and the secondary treatment of the wastewater products intended to be
recycled to agriculture.
For the primary treatments, the effects of handling the urine and the faeces were
compared, i.e. urine-diversion and faecal separation were compared to treatment at
a large wastewater treatment plant (WWTP). When evaluating the quality of the
fractions other, in the wastewater treatment plant, co-treated materials were
considered as contaminants, i.e. greywater and industrial wastewater. As the
phosphorus in the greywater also is recycled in the sludge from the wastewater
treatment plant it was included as a recyclable nutrient. The different systems were
compared according to Priority Group 1 (Kärrman & Jönsson, 2000) the amount of
N, P and K it was possible to recycle to the arable land, the amount of heavy
metals contaminating the recycled fertilisers, water emissions and according to
Priority Group 2 (Kärrman & Jönsson, 2000) the total energy usage for the
systems. Other factors in Priority Group 2 (Kärrman & Jönsson, 2000) were not
included in the comparison.
For the secondary treatments, the comparison was performed regarding
disinfection efficiency; risk for regrowth; and energy usage. The different
treatment scenarios compared were storage; liquid composting; dry composting;
pasteurisation; and chemical treatment by urea and peracetic acid (PAA).
Comparison of different primary treatments
The comparison between the different systems was expanded so the benefits from
the systems according to fertiliser recovery were equal (Figure 19). When
differences in production of nutrients between the systems investigated were
identified, the most favourable system was set as the standard and the other
systems were charged with the environmental effects from external production of
nutrients to arable land (Figure 19). No compensation for differences in heat or
biogas production was included in the comparison. One of the main heat sources
was the treated wastewater itself, from which the heat could be recovered by a heat
exchanger, but as no change in the water consumption was taken into account, no
differences in this heat recovery were noted between the different treatment
systems. The change in biogas production could not easily be foreseen in the
different treatment scenarios, and therefore the biogas production in the sewage
treatment plant was not included in the comparison.
Figure 19. The main system and the extended system in the comparison for the different
wastewater treatment systems. In the evaluation the toilet water treated and the amount of
nutrients (N, P and K) to arable land are kept equal for the different systems.
As the greywater has to be treated in all of the systems discussed, it was not
included as an energy consumer during the treatment, except for the extra energy
needed when calculating the production of the extra phosphorus fertilisers to equal
the large wastewater treatment plant system as the phosphorus in the greywater is
included in the recycled sludge. One alternative for treatment of the wastewater,
especially as an alternative for single households, is to use a septic tank combined
with a soil filter. The energy usage in such a system is minimal but so is the
nutrient recovery and, as the production of the nutrients is the main energy
consumer and none of the nutrient resources are recovered, this kind of system was
not included in the comparison.
Nutrient recycling
Depending on collection strategy for faeces and urine, different amounts of
nutrients can be collected. If a blackwater (BW) system is used, all of the nutrients
from the urine and the faeces can be collected. This corresponds to approximately
90% of the N, 74% of the P and 79% of the K in household sewage water (Figure
20). However, these systems usually have problems with too large amounts of
water diluting the collected fraction, both large amounts of flushwater and other
water added to the toilet, such as domestic cleaning water. When the toilet waste
was collected separately, no sludge from the sewage treatment plant was assumed
to be recirculated to agricultural land, nor was the handling of this sludge included
in the evaluation of these systems.
Recyclable amounts (%)
UD 100%
UD 70%
UD 100% FS 85%
UD 70% FS 85%
UD 70% FS 58%
Figure 20. Amounts of nutrients from the household wastewater collected and available for
recycling to arable land depending on collection strategy or treatment method,
BW=Blackwater UD=Urine-diversion FS=Faecal separation, WWTP=sludge from a large
wastewater treatment plant, the following percentage figure gives the amount of the fraction
When only the urine is collected, by using urine-diverting toilets, and all of the
urine is collected (UD 100%) 79%, 49% and 58% of the N, P and K, respectively
are collected. When, as in Ekoporten (Paper III), 70% of the urine was collected
(UD 70%) the collected amounts were 55% N, 35% P and 40% K (Figure 20). A
combination of well motivated users and well functioning urine-diverting systems
could increase the collection of the urine to closer to 100% diverted than the 70%
measured in Ekoporten (Paper III).
When urine-diversion is combined with faecal separation (FS) more nutrients are
collected. The optimal system would be if all the urine was collected and, as in the
pilot-scale study, 85% of the faecal nutrients were separated. Then 88% of the N,
71% of the P and 76% of the K would be collected in a low volume easy
recyclable fraction (Figure 20). If the system were more like the one in Ekoporten,
where 70% of the urine was diverted and on average 58% of the faecal nutrients
were separated, 62% N, 49% P and 53% K would still be separated to a low
volume and easily recyclable fraction (Figure 20)
Quality of the recycled products
When comparing the collectable and recyclable amounts of the nutrients from the
wastewater to the possibilities for nutrient recycling via the sewage sludge from
wastewater treatment plants, urine-diversion and faecal separation recycle less P,
as the greywater contributes 26% of the P in the wastewater, and significantly
more N and K are recycled (Figure 20). An additional effect of separate collection
of the toilet waste is the avoidance of heavy metals and other contaminants from
greywater and other sources (Paper I). These normally pollute the sewage water
and finally end up in the sludge (Table 12). So when aiming for recycling of
unpolluted nutrients from households, a large proportion of this can be achieved by
recycling the separated toilet fractions.
Table 12. Amounts of heavy metals in mg per person and year that accompany the
recyclable nutrients in Figure 20 BW=Blackwater UD=Urine-diversion FS=Faecal
separation (Paper I), WWTP=Sewage from a large-scale wastewater treatment plant
(Balmér, 2001)
Cu Cr Ni Zn Pb Cd Hg
BW 440 11 30 3900 8.0 4.0 3.6
UD 100% 37 3.7 2.6 16 0.73 0.25 0.3
UD 100% FS 85% 380 9.9 26 3300 6.9 3.4 3.1
UD 70% FS 58% 260 6.8 17 2300 4.7 2.3 2.1
WWTP 12000 1400 2100 22000 1600 56 36
If the urine and the faeces are removed from the sewage water, only the greywater
is left. This is a large volume fraction with a relatively high total amount of
contaminants, both heavy metals and organic pollutants, which have to be treated.
Most sewage treatment systems of today are not constructed for the composition of
the greywater, and in some cases industrial water and stormwater, but rather for
mixed household wastewater. The focus for sewage treatment during the past
century has been the large environmental effects such as polluting particles and
eutrophication by phosphorus discharge. Therefore, treatment plants are currently
adapted to the functions of removing these fractions from the wastewater. If the
main proportion of the nutrients is to be removed, the purpose of the sewage
treatment plant will have to be rethought, in particular regarding how their
function may have to be adapted to remove the pollutants such as heavy metals
and organic substances to prevent them from entering the environment. Thus the
focus will have to be kept more on protecting recipient waters and less on how to
make the sewage sludge as clean as possible. Effects of the organic pollutants in
effluent have already been identified, e.g. changes in the sex of male fish caused
by organic hormone mimicking substances in wastewater (Tilton et al., 2002).
Water emissions
If the remaining wastewater after some or all of the toilet water is removed are
treated in a large wastewater treatment plant the emissions of phosphorus will not
be affected as the emissions are regulated as concentrations in the effluent.
However, the need for treatment and precipitation chemicals will decrease as will
also the energy usage for nitrogen removal. The major effect will be for the
emissions of potassium, as the regular treatment only removes a few percent of the
potassium, and for nitrogen in systems where the nitrogen not is removed.
If the same system of sewage treatment is used after the majority of the nutrients
have been removed by urine-diversion and faecal separation (Figure 20), the need
for chemical precipitation of phosphorus will decrease and if no changes are made
to the treatment system.
Energy usage
Looking at the total energy usage of water production and wastewater treatment,
water production uses approximately 40% of the energy (Kärrman et al., 1999;
Kärrman, 2000). The amount of energy used in water production is mainly
correlated to the volume produced. Therefore, differences in water consumption of
different toilet options have a significant impact on total energy usage. However,
this was not taken into account, as water consumption is hard to predict for the
different toilet systems. When used correctly, the water-saving potential of urine-
diverting toilets is high, especially for those where the faeces are collected dry, e.g.
in Gebers only approximately 0.5 m3 of flushwater was used per person and year
(Andersson & Jensen, 2002) compared to the Swedish design valies of over 18 m3
p-1 y-1.
The energy usage for treatment of the toilet water in the wastewater treatment
plant was calculated according to the flow and the composition of the incoming
wastewater. Used energy for the treated water volume was calculated as the energy
for pumping, in the sewers (0.1 MJ m-1) and in the wastewater treatment plant (2.7
MJ m-1) in Gothenburg (Balmér et al., 2002). For the removal of BOD7 and
nitrogen the aeration energy were according to Balmér et al. (2002) 3.6 MJ kg-1
and 10.3 MJ kg-1, respectively (Table 13). The energy for producing chemicals for
phosphorus removal (Frodhagen, 1997) was included in the energy usage of the
wastewater treatment plant in Table 14. Approximately 20 kg of PIX 111 was
assumed to be used per kilogram of removed phosphorus (Tidåker, 2002), which
corresponded to an energy usage of 55 MJ kg P-1 (Table 13), and the amount of
phosphorus removed corresponded to the average percentage removed in the
sewage treatment plant Henriksdal in Stockholm (98%) (SVAB, 2002). These
values were then used for calculating the energy needed for treatment of the water
in a large-scale wastewater treatment plant.
Table 13. Energy usage for treatment of wastewater in a regular wastewater treatment
plant (Balmér et al., 2002) and for production of PIX 111 for phosphorus removal
(Frodhagen, 1997)
Variable Unit Energy usage
Flow MJ m-3 0.37
Biological oxygen consumption MJ kg-1 BOD7 3.6
Nitrogen MJ kg-1 N 10.3
Production of PIX 111 for P removal MJ kg-1 P 55
In the large wastewater treatment plant only the energy usage for treating the toilet
water was estimated, but due to the uncertain water usage by the different systems,
the same amount of treatable flushwater was used for the different systems
evaluated except for the blackwater, as no water enters the sewer in this system
(Table 14).
Table 14. The energy usage of electricity, steam and, oil (MJ p-1 y-1) in the different steps of
wastewater treatment in the systems evaluated and the corresponding transport length (km,
one way) possible using an ordinary truck (Sonesson, 1998), and with the same total energy
usage in all systems. The systems compared were: blackwater collection (BW); urine-
diversion (UD) with different diversion efficiency (in percent); and in some cases combined
with faecal separation (FS) with different separation efficiency (in percent); and finally a
large sewage treatment plant (WWTP)
Alternative WWTP Nutrients Spreading Total Transport
Sort MJ p-1 y-1 MJ p-1 y-1 MJ p-1 y-1 MJ p-1 y-1 Km
BW 0 5.4 13 18 117
UD 100% 27 42 12 81 106
UD 70% 47 109 8.7 165 91
UD 100% FS 85% 9.7 11 13 34 113
UD 70% FS 85% 30 79 9.5 118 103
UD 70% FS 58% 36 88 9.2 133 100
WWTP 95 192 5 292 0
The energy usage for production of supplementary nutrients, which was needed to
equalize the amounts of nutrients in all systems. The amounts of recycled
household wastewater nutrients were set to be: 90% of the nitrogen as in the
blackwater, 98% of phosphorus as in the WWTP scenario and 79% of potassium
as in the blackwater (Figure 20). The energy values for production of
supplementary nutrients were taken from Davis and Haglund (1999), and urea
(47% N) (52 MJ kg N-1) was used as the mineral nitrogen fertiliser due to its
similar plant availability compared to the nitrogen recirculated from the toilet
fractions, P2O5 (21% P) (31 MJ kg P-1) was used as the phosphorus fertiliser and
mined KCl (52% K) (5.7 MJ kg K-1) was used as the potassium fertiliser. The
energy usage for the fertiliser production is displayed in Table 14 as the sum of the
total energy usage from electricity, steam and, oil.
The energy usage for the recycling of the plant nutrients was calculated based on
spreading of the liquid material with a hose spreader at a dosage of 80 kg plant
available nitrogen per ha. The mass spread per person in the blackwater system
was assumed to be 550 kg of urine plus 550 kg of flushwater and 195 kg of
separated faecal matter (19.5 kg of dry matter with a dry matter content of 10%),
and the separated mass was assumed to correspond to the amount of separated
nutrients. Approximately 9-13 MJ was used per person and year for spreading the
urine and the faeces in the separating systems. In the other sorting systems (urine-
diversion and faecal separation) the percentage not collected were deducted from
the mass and the nutrient content of the material spread. The sewage sludge was
assumed to have a dry matter content of 20% and therefore be spread as dry matter
using approximately 5 MJ p-1 y-1, the dry matter mass collected was set to 60 g per
person and day (Balmér, 2001) in the WWTP system.
The main energy usage was for the supplementary production of nutrients to bring
the same amount of plant available nutrients to arable land in all cases (Table 14).
Summing up all the energy used, the conventional wastewater treatment (WWTP)
used the most energy (Table 14). In the other cases, the energy saved was used to
calculate the possible distance for transporting the material by an ordinary truck
(Sonesson, 1998) without exceeding the energy usage by the WWTP system.
For a system like Ekoporten, where 70% of the urine was diverted and 58% of the
faecal nutrients were separated, the maximum distance for transporting the
material was 97 km, while the blackwater system permitted up to 116 km of
transport without using extra energy (Table 14). The most favourable system
combining both urine-diversion (100%) and faecal separation (85%) allowed a
transport of the material up to 111 km, with an ordinary truck (Sonesson, 1998),
before spreading.
Comparison of secondary treatments to attaining high hygiene
Comparing the different methods available for secondary treatment of the
wastewater products before recycling, many different indicators can be used to
evaluate the efficiency. However the most important of these is the reduction of
pathogens to produce a hygienically safe material.
The treatment methods compared here were: storage; pasteurisation; liquid
composting; dry composting; urea treatment; and treatment by peracetic acid
(PAA). The emphasis of the comparison was whether the treatment could assure a
safe waste product; the scale; and the energy usage.
The simplest secondary treatment is storage. However, storage cannot guarantee a
product considered to be safe according to hygiene standards unless several
external conditions such as temperature, time and composition of the material,
especially the ammonia content, are kept at predetermined levels. These factors
affect the final product and changes in them mean that storage does not inactivate
all organisms, especially those that have some kind of dormant state that can
provide them with protection for several years. In the chemical treatment of faecal
matter, water was added as the control (Paper V). In that study the amount of E.
coli and faecal Streptococci spp decreased during the treatment. No E. coli was
detected after 50 days, which corresponds to a reduction of >6 log10. Salmonella
that was added to approximately the same initial concentration was still detectable
after 50 days of treatment. An initial increase of Enterococcus spp was determined
in the control (Paper V), but after 50 days the presence of Enterococcus spp had
decreased to approximately 10% of the initial amount. No significant difference
was detected for the number of Salmonella typhimurium 28B phage. However, the
Ascaris suum eggs added to the material showed a significant reduction in viability
after 50 days of storage of faecal matter, approximately 0.2% viability compared
to 90% in the control (Paper V). Storage can however also give large
environmental effects. The organic matter is degraded and in some cases results in
waste products with negative environmental effects, such as methane from
anaerobic conditions and nitrous oxide from incomplete nitrification or
denitrification. Both gases are potent greenhouse gases. If all the sewage sludge
produced in Sweden were to be stored, the total national emissions of methane
would increase by 0.2% and the nitrous oxide would increase by 5% (Flodman,
2002). The advantage with storage is that it can be performed in whatever scale
needed and the extra need for energy is minimal. Still this is not a preferable
method for treatment of sewage products, except urine, which has an additional
disinfection effect sufficient for inactivation of pathogens from the urea content,
similarly to chemical disinfection by added urea (Höglund, 2001). The collected
urine mixture also contains only small amounts of organic matter that can be
degraded. No hydrogen sulphide has been detected in urine storage tanks
indicating low or no anaerobic activity in urine mixture.
Composting of sewage products can either be performed at a dry matter content of
2-10%, liquid composting, or at approximately 35% dry matter, dry composting.
Liquid composting is a complex technique as it is performed in a reactor where
air is forced into the liquid. Therefore it is also quite energy demanding;
approximately 140 MJ per person and year are needed to treat blackwater (1295 kg
p-1 y
-1) in a 17.5 m3 reactor (Skjelhaugen, 1999b). If only the separated faecal
water (195 kg p-1 y-1) is treated, the energy demand corresponds to 18 MJ p-1 y-1
(Skjelhaugen, 1999b). The material in the compost reactor is liquid and the process
is fully mixed, so disinfection is achieved in all of the material. The reactor
facilitates monitoring of both incoming and outgoing air, which makes it possible
to preheat the incoming air to avoid low temperature areas and to condense the
outgoing air. Condensing the outgoing gas makes it possible to collect a major part
of the components in the gas and then percolate them back into the reactor. Further
treatment of the air can be done in a bio-filter (Haug, 1993). It is also possible to
collect the energy in the outgoing air, as the compost reaction is exothermic and an
energy surplus is produced. The main use for this energy is to preheat the
incoming air. The reactor also makes it possible to monitor the disinfection, as all
of the material attains the same temperature and the equation in Table 10 can be
used to determine the required time of treatment according to the required safety
margin. A disadvantage of liquid composting is that the highly technical reactor
needs to be large to be an economically viable alternative and that it uses large
quantities of electrical energy.
Dry composting uses the same biological mechanisms as liquid composting but
the higher dry matter content makes natural ventilation possible as the thermal
movement in the material gives enough oxygen if the structure is porous. Dry
composting can either be performed in reactors or in open windrows. The
treatment is not as scale-dependent as liquid composting since small-scale local
composts can be used, although the smaller the scale, the better the insulation has
to be. In small-scale systems it is often impractical to have pre-heating of the
incoming air and treatment of the outgoing air. Therefore cooled zones will occur
around the air inlet and the outgoing air can have effects on the environment such
as smell and ammonia emissions. However, a lot of the emissions will probably
condense on the lid of the composting reactor and percolate back into the material.
The cool zones can be handled by having material with a better hygiene standard
close to the air inlet, e.g. food waste, which can be used as a pre-heater. One major
disadvantage with the composting of sewage products is that it requires handling
of the raw untreated matter when starting and turning the compost. The
environmental effects from the treatment are small if the outgoing air is
condensed. There is also a risk for recontamination of the material after the
Pasteurisation is an efficient treatment method for disinfection. The duration of
treatment and treatment temperature can vary according to the equations in Table
10 depending on the safety margins used. The system for thermal treatment is
energy demanding and highly technical, and is therefore only an alternative for
large-scale treatment. The energy needed to get temperatures as high as those in
the composts (65°C), on average increasing the temperature by 55°C, would be
39 MJ p-1 y
-1 for treatment of the separated faecal matter, while to treat the
blackwater (1295 kg p-1 y
-1) would require 300 MJ. By using a heat exchanger
recovering the energy of the treated matter, it would probably be possible to reuse
half of the energy to approximately the amount used by liquid composting. The
treated matter is hygienically safe but the risk for recontamination is large as all
organic matter is more or less unaffected and the initial competition from other
organisms is small as most of them are killed during the treatment.
The treatment with urea only requires that the treatment is performed in a closed
container where air exchange is low, as this prevents the ammonia produced from
being lost. The energy usage for production of the urea added according to the
dosage in Paper V (30 g of urea nitrogen per kg of material to treat) for treatment
of separated faecal matter (195 kg p-1 y
-1) with a dry matter content of 10%
corresponds to 150 MJ p-1 y-1. However, if the disinfected material is recycled to
agriculture the energy usage for production of this urea can be fully allocated to
the fertiliser, as all of the urea used acts as a fertiliser upon application to land. The
increased nitrogen content somewhat increases the energy usage as the driving
distance for spreading the material increases. However, the total energy usage for
spreading is relatively small compared to the energy used in the production of
fertilisers and can therefore be disregarded. Furthermore, the dosage can probably
be decreased significantly, which will be tested in future studies. As long as the
ammonia is still in the treated matter there will be no risk for recontamination. As
the only requirement is that the air is not ventilated, emitting the ammonia, there is
no limitation on the size of the treatment. The only proven efficient treatment
temperature so far is at 20°C. According to Allevi et al. (1994) and a temperature
of at least 10°C is needed for inactivation by ammonia. Therefore urea treatment
should be performed at temperatures above 10°C until tests have shown
satisfactory effects at lower temperatures.
The PAA treatment, like the urea treatment, is not restricted in size. However,
proper mixing of the material is necessary, as the rapid treatment requires a totally
mixed material. Given this condition, the main treatment effect can be achieved
within less than 12 hours of treatment. As the chemicals are strongly oxidising, a
higher organic matter content in the material requires an increased dosage of
chemicals. The energy for producing the chemicals for disinfection of the faecal
matter (195 kg p-1 y-1) is approximately 20 MJ, if the urine were included; the need
for chemicals would probably not increase particularly. As the reaction is fast and
the active substance is consumed during treatment, the treated matter has to be
applied to land soon after treatment, otherwise there is a large risk for re-infection
(Paper V).
Concluding summary
As regards the general composition of household wastewater, the urine mixture
contains the major proportion of the nutrients and only a small fraction of the
heavy metals (Paper I). Compared to the average composition of the wastewater,
the faecal matter is also a relatively clean fraction. The faecal fraction contains
smaller amounts of nutrients and larger amounts of heavy metals compared to the
urine but it has only one tenth of the mass of the urine. These two wastewater
fractions are most interesting for recycling due to their nutrient content, their
quality of the nutrients and their relatively small volumes (Paper I).
Due to the origin of these fractions, they are initially easy to collect without
requiring large amounts of energy. To get the same amount of unpolluted nutrients
from mixed wastewater requires a lot of energy and highly technical systems. If
the urine is collected separately in a urine-diverting toilet system, there is only an
increase in energy usage for transporting the urine to the farms and spreading it.
The faecal matter can either be collected dry in a dry urine-diverting toilet or
separated from the flushwater of a double flush urine-diverting toilet, after a short
pipe transport, in an whirlpool, surface tension separator as the faecal nutrients are
particle-bound (Paper II, III).
In investigations of different urine-diverting sewage systems, between 55% and
90% of the urine-derived nutrients have been recovered (Jönsson et al., 1997,
1998, 1999, 2000; Vinnerås, 1998; Lindgren, 1999; Johansson et al., 2000;
Andersson & Jensen, 2002). A larger proportion of the urine nutrients is probably
collected in more recently built systems, as toilet design has developed and the
knowledge of why the system is used and how it functions has increased. If the
urine is collected and recycled, the energy saved compared to treating it in a large
sewage treatment plant and producing the corresponding amount of mineral
fertilisers would correspond to a truck-driven transport of the urine mixture of
approximately 104 km one way.
One way to collect both the urine and the faecal matter is by using low flush,
mainly vacuum, toilets. An alternative for recovering the faecal fraction while still
having a flush toilet is to separate the faecal matter from the flushwater after a
short pipe transport. To work, this requires that the faecal nutrients remain in the
faecal particles (Paper II) and a urine-diverting toilet has to be used. As the faecal
particles are rapidly degraded in the water the separation has to be performed
within a short period of time after the flush (Paper II). By using a whirlpool,
surface tension separator, it was possible to collect up to 85% of the faecal
nutrients, and the dry matter content of the collected matter was approximately
10% when toilet paper was included.
In a well functioning whirlpool, surface tension separation system, misplaced
wastewater such as the domestic cleaning water in Ekoporten (Paper III) only has
small effects on the content of non-faecal matter, such as heavy metals, in the
separated solids. Thus this kind of system seems be robust even if the toilet is
misused to a small extent as the water soluble and suspended substances are not
separated into the separated solids fraction.
In the final comparison of the primary wastewater treatment the wastewater
treatment plant was set as the normative system (Table 15). When comparing with
the other systems, if the system was more than 50% better or worse than the
reference it was given as + or -. If the difference were less than 50% it was given
as 0.
Table 15. Comparison between different primary treatment systems for treatment of
household wastewater. The compared systems were: blackwater collection (BW); urine-
diversion (UD) combined with faecal separation (FS) with different separation efficiency
(in percent) and finally a large sewage treatment plant (WWTP)
BW UD 100%
FS 85%
UD 70%
FS 58%
N, P & K to land + + + 0
Heavy metals to land + + + 0
N, P & K to water + + + 0
Metals to water 0 0 0 0
Energy usage +* +* +* 0
Local transport - - - 0
* the treatment of the collected blackwater was not included in the energy usage
The systems where the toilet water was collected separately were significantly
better regarding the recycling of unpolluted nutrients, nutrient losses to the
recipient waters and decreasing the energy usage compared to the conventional
sewage treatment in a large sewage treatment plant (Table 15). As the content of
heavy metals in the toilet water was significantly lower than the total flow to the
sewage treatment plant the amounts of heavy metals to land will be lower in the
systems where the toilet fractions is recycled separately.
When comparing the different systems where the urine and faeces are collected
separately the blackwater system will show the best results according to its energy
usage and environmental effects. However, today there is no reliable and well-
functioning toilet system available where the blackwater is collected with a high
dry matter content >4%. One of the main problems is the large water volume that
dilutes the fraction and increases the volume to be handled. One method to keep
the water volume low is by use of vacuum toilets especially urine diverted ones.
These systems are technical and sensitive towards malfunction and if the vacuum
goes down no toilet will be possible to flush. Also material poured into the toilet,
e.g. domestic cleaning water, will be collected together with the blackwater
increasing the volume and maybe also pollute the otherwise relatively unpolluted
urine and faeces. The second best system according to the compared factors (Table
15) is the collection of the urine and the faeces via urine-diversion and faecal
separation. However, it is not possible to collect the same large amounts of
nutrients and thereby not possible to have as large positive environmental effects.
But if the urine-diversion is done correctly most of the urine should be possible to
collect, and the faecal nutrient seems to be possible to collect up to 85% and
maybe even more by further development of the system. One other major
advantage with the urine-diversion and faecal separation system is its robustness
due to the absence of moving parts and the small effects of misusage on the
function, except for the loss of nutrients. The system with faecal separation will
also be possible to use in areas where water instead of paper is used for anal
cleansing as almost no water will be found in the fraction of separated solids when
no particles are flushed.
The separation of the faeces can be performed in several ways. One very important
factor for the separation is that the separated particles are removed from the water
directly after the separation. Otherwise the nutrients will be dissolved into the