3. Organic Pollutants
3.
Organic Pollutants: sources, pathways, and fate through urban
wastewater treatment systems
3.1
Sources and pathways of organic pollutants in UWW
There are a large number of organic pollutants from a wide range of sources which may
enter UWW. Paxéus (1996a) identified over 137 organic compounds in the influent of the
municipal wastewater plants in Stockholm. The physical and chemical properties of some of
these organic pollutants are outlined in Appendix B. The main categories of organic
pollutants detailed in this report are:
Polycyclic Aromatic Hydrocarbons: Polycyclic aromatic hydrocarbons (PAHs) arise from
incomplete combustion or pyrolysis of organic substances such as wood, carbon or mineral
oil. Such combustion processes include food preparation in households and food shops;
discharge of certain petroleum products (from garages, vehicle washing and maintenance,
fuel stations); discharge of storm runoff with PAHs from car exhaust particles and road
runoff; and also from incomplete combustion processes in urban landfills.
The most frequent anthropogenic sources of PAHs are: house fires, heat and energy power
stations, vehicle traffic, waste incineration and industrial plants (cement works, metal
smelting, aluminium production). Forest fires represent natural sources. PAHs concentrate in
sewage sludge due to their low biodegradability.
Polychlorinated Biphenyls (PCBs):
There are two main sources of PCBs:
• Directly manufactured PCBs (by chlorination of biphenyls), used as hydraulic liquids
(hydraulic oils), emollients for synthetic materials, lubricants, impregnating agents for
wood and paper, flame protective substances, carrier substances for insecticides and
in transformers and condensers. The EU1996 PCB Disposal Directive 96/59/EC
requires the phasing out of all PCBs by 2010 or by 1999 under international
agreement by the North Sea States. Existing transformers and other electrical
equipment which contain 50-500 mg.kg-1 PCB may be retained in service until the
end of their useful life.
• The other main source of PCBs in the environment are combustion processes, from
waste incineration plants, fossil fuel burning and to other incomplete combustion
processes.
PCBs are adsorbed by solids and therefore they accumulate in sewage sludge. The highly
substituted (high chlorine content) PCBs are the main representatives potentially present in
sewage sludge, while they amount to just 35% of the total technical PCBs. Recycling of
PCBs in the environment is very important and remediating historical pollution would be
necessary if the background levels found are to be reduced.
Di-(2-ethyhexyl)phthalate (DEHP):
DEHP is used as emollient in synthetic materials. In Germany, 90 % of DEHP is used in
PVC and about 10% in laquers and paints. It is common to use DEHP as antifoaming agent
in paper production, as an emulsifier for cosmetics, in perfumes and pesticides, they aid in
the production of different synthetic materials such as dielectric in condensers, and
substitute for substances such as PCBs and pump oil. DEHP specific emissions from
various human activities have been identified by Bürgermann [1988] as follows:
•
•
•
•
•
•
cellulose/paper production
DEHP production
plastisol-coating process
PVC production and processing, leaching from PVC products
leaching from waste in landfills
waste incineration and uncontrolled combustion
64
3. Organic Pollutants
DEHP is found regularly in municipal wastewater and, because of its lipophilic properties, it
concentrates in sewage sludge.
Anionic and Non-ionic Surfactants:
Surfactants are contained as the main active agents in all washing and cleaning agents.
These compounds are covered in detail in Case Study (f).
Polychlorinated Dibenzo-p-dioxins and Dibenzofurans (PCDD/PCDF):
The generic term "dioxins" represents a mixture of 219 different polychlorinated dibenzo-pdioxins and furans. The most well known and hazardous dioxin, is the tetrachlorodibenzo-pdioxin (TCDD). Dioxin concentrations are calculated as sum of the toxicity equivalents (TEQ)
relevative to the most toxic dioxin [TCDD].
The three main sources of polychlorinated dibenzo-p-dioxins and dibenzofurans are as
follows [Mahnke, 1997, Horstmann, 1995]:
• Chemical reactions or chemical reaction processes: Dioxins arise as unwanted
by-products from the production or use of many organo-chlorine compounds, such as
chlorine bleaching of cellulose in paper production and chlorine alkali electrolysis. In
these cases the formation mechanism can be explained by substitution,
condensation or cyclisation reactions.
• Combustion processes or thermal processes: Dioxins arise by thermal processes
and are released into the atmosphere. The dioxin formation results from a de-novo
synthesis. Important thermal sources are:
o waste incineration plants and incomplete combustion processes in landfills;
o combustion plants;
o iron smelting;
o sinter plants, non-ferrous smelting and recycling plants;
o petrol and diesel engines.
• Dioxins can also arise from all incomplete combustion processes involving
chlorine. This explains the ubiquitous dioxins occurrence in the environment.
Anthropic production of dioxins has predominated since the introduction of
organochlorine compounds in industrial applications (1920). With the improvement of
the catalysts in waste incineration plants and other measures for reducing the dioxins
emission, the fraction of anthropic dioxins has been declining since 1970. Dioxins
can be formed and released into the atmosphere also by natural events, e.g. forest
fires. Dioxins can also be generated by the biochemical transformation of precursor
compounds (for example during degradation of chlorophenols).
Dioxins speciation in household wastewater and laundry wastewater is similar to those in the
sediments of UWW collecting systems and sewage sludge. A mass balance indicates that 27 times more dioxins in sewage sludge originates from households than from urban runoff.
Washing machine effluent is a major source of dioxins in household wastewater. Dioxins
were also detected in shower water, and in urban run-off from various human activities
[Horstmann 1993, 1995]. These results suggest that the importance of household
wastewater as a dioxin source has been underestimated [Horstmann et.al., 1993,
Horstmann, 1995].
Sources of other potential organic pollutants are listed below:
Organic pollutants can originate from food and household related products, such as long
chain fatty acids and their methyl and ethyl esters, originating from faeces, soaps and food
oils. Being relatively hydrophobic these compounds are attached to particles, the
concentration of fatty acids and esters in the unfiltered influent is more than 500 µg/l. Other
organic pollutants from domestic origin are the sterols from animal foods and faeces and
indol from faeces. Caffeine is also found from discharges from coffee processing.
65
3. Organic Pollutants
Plasticisers and flame retardants are still used in many products for household and
industrial applications. Among the organic pollutants present are benzenesulphonamides,
adipates (esthers of hexandioic acid), phthalates (esters of phthalic acid, among which
DEHP is the most common), and several phosphate esters. (2-chloroethanol phosphate) and
TBP (tri-n-butyl phosphate) are used in flame-retardant compositions in textiles, plastics as
well as in other products.
Preservatives and antioxidants are constituents of household and industrial products, and
among the organic pollutants linked with these compounds are parabens (esters of
hydroxybenzoic acid), and also substituted phenols and quinones are among the
constituents.
Solvents both chlorinated and non-chlorinated (alcohols, ethers, ketones) are present in a
large range of products such as car shampoos and degreasing products, household
cleaners and degreasing agents from vehicle maintenance and production. Chlorinated
solvents, such as trichloroethylene and trichloroethane, are in increasingly wide use: the
amounts consumed in France per year are 24,000 and 28,000 tonnes, respectively. The
principal sources of diffuse pollution from chlorinated solvents are due to artisanal activities
such as metal finishing activities and dry cleaners. Nevertheless, domestic sources from
aerosols and other agents are not negligible. Pollution by metal cleaning activities is usually
considered as diffuse discharges as they are usually from small firms with only few
employees. Garages consumed around 15,000 tonnes of solvents in 1988, about 60% of
which is lost to the atmosphere and the rest as waste. Of the 6,000 tonnes, of waste solvent
some will be discharged into the UWW collecting system [Agences de l'Eau, 1993]. Metal
finishing used 50,000 tonnes of solvents in 1991 and their aqueous wastes are discharged
into UWW collecting systems, although these are usually in low levels. Dry cleaning
consumed around 19,500 tonnes of solvents in 1988 and it has been determined that
0.3x10-3 kg of solvent/100 kg of clothes cleaned ended up in wastewater.
Fragrances from households, beauticians and hairdressers, generate mixtures of terpenes
and synthetic musks (galaxolides), and are also found in industrial detergents. These are
covered in more detail in Case Study E, Section 6.
Pesticides and herbicides are also a common component of the urban wastewaters and
they result from road and rail weed treatment, and from gardens, parks and urban woodland
areas. They include the triazine group, the phenyl urea group (e.g. chlorotoluron, isoproturon
and diuron), the phenoxy acid group (eg. Mecoprop and 2,4-D) and glyphosate [Revitt et al.,
1999].
An enormous quantitiy of pharmaceutical products are prescribed every year: 100 tonnes
of human drugs were prescribed in 1995 in Germany [Ternes, 1998]. Pharmaceuticals in the
Urban Environment are discussed in Case Study (d), Section 6.
Triclosan (2,4,4’trichloro-2’hydroxydiphenyl ester) has been used in soaps, shampoo and
fabrics, as an antimicrobial agent. While these compunds are regarded as low toxicity their
2-hydroxy isomers have been shown to undergo thermal and photochemical ring closure to
form polychlorinated dibenzo-p-dioxins which are highly toxic. (Okumura et al 1995).
66
3. Organic Pollutants
3.1.1 Domestic and Commercial Sources
A study carried out in France in 1995 by ADEME, showed the sources of the main organic
micropollutants in sludge from WWTS were mainly domestic and commercially related (see
Table 3.1). Another study, by SFT (in collaboration with the wider Norwegian government
environmental study programme and the A/S Sentralrenseanlegget RA-2 WWTS),
investigated sources of PAH, PCB, phthalates, LAS and NPE. This study found that sewage
from domestic sources, in this instance from an isolated housing estate with a separate
sewage and stormwater drainage system, does make a significant contribution of the above
organic pollutants to urban wastewater [SFT report 98/43].
Table 3.1 Principal sources of organic micropollutants in urban wastewater treatment
works [ADEME, 1995] +++ very likely, ++ likely, + less likely present
POLLUTANT
ORIGIN
Domestic
usage
Storm
runoff
Commercial
effluent
Aliphatic
hydrocarbons
Monocyclic aromatic
hydrocarbons
PAHs
Fuel
++
++
++
Solvents, phenols
+
+
++
By-products of petrol
transformation and
insecticides
Solvents, plastics,
chlorination
Solvents, pesticides
+
+
+
++
+
++
+
+
++
(+)
+
+
+
+
+
++
++
++
++
0
+
+
++
++
Halogens
Chlorophenols and
Chlorobenzenes
Chlorinated PAHs
Pesticides
Phthalate esters
Detergents
Nitrosamines
PCB, hydraulic fluids
Plastifier
Industrial by-products
(rubber)
Soil is also a major repository of organic matter and the soluble fractions can leach/run-off in
to water courses, especially in upland areas where measures to remove colour and
formation of trihalomethanes during drinking water treament is important.
A.
PAHs and PCBs
Table 3.2 shows that the PAH concentration profiles for three Swedish WWTS varies. This
may in part be due to differences in the catchment areas, with the sources of the pollutants
coming from different local industries. Most of these PAHs are expected to derive from
diffuse commercial activities and traffic but PAHs such as pyrene, which is believed to be
derived from at least 50% domestic sources, is present in all the samples at more consistent
concentrations than some of the other compounds.
Mattson et al (1991) referenced in Paxéus (1996a) found that PAHs from food, an often
overlooked source of this pollutant, from households can reach 50-60 % of the total UWW
collecting system load for pyrene and phenanthrene. This is an important observation as
household sources of PAHs are likely to be more difficult to control than commercial
sources.
Another source of PAHs from domestic and commercial activities is the use of phenol and
creosol in products such as wood preservatives. In Finland, 430 tonnes of wood
preservatives were used in 1995 [Finnish Environmental Institute, 1997]. PAHs may enter
UWW as a result of spillages or as surface runoff from rainwater.
67
3. Organic Pollutants
Table 3.2 PAHs concentrations in urban waste waters in Sweden [Paxéus 1996a]
PAHs
Naphthalene, dimethyl
Naphthalene, methylpropyl
1,1’- Biphenyl, dimethyl
1,1’- Biphenyl, ethyl
Anthracene/Phenanthrene
Methyldibenzothiophene
2,8-Dimethyldibenzothioprene
Anthracene/Phenanthrene methyl (different
isomers)
Anthracene/Phenanthrene dimethyl (different
isomers)
Retene
Pyrene
Pyrene, methyl (different isomers)
Pyrene, methyl, methylethyl or tetramethyl
1,1-Diphenylethane
HST
µg/l
1
3
2
1
1
<LOD
5
2
WWTS
GRYAAB
µg/l
0.5
<LOD
0.5
<LOD
<LOD
<LOD
<LOD
<LOD
SSW
µg/l
<LOD
<LOD
<LOD
<LOD
0.5
0.5
<LOD
3
1
<LOD
1
<LOD
3
2
1
<LOD
<LOD
<0.5
<LOD
<LOD
0.5
0.5
2.5
1
1
<LOD
1, (H)- Indene, 1-phenylmethylene
0.5
<LOD
<LOD
9H-Flouren-9-one
0.5
<LOD
<LOD
2-Anthracenaemine
<LOD
<LOD
9
Acridine, 9-methyl-Dibenz(b,f) azepine
<LOD
<LOD
0.5
Octahydrophenanthrene, dimethyl-, isopropyl
0.5
<LOD
<LOD
Total PAH
23.5
<2
19.5
HST = Henriksdal Sewage Treatment Plant, GRYAAB = Gothenburg Regional Sewage
Works, SSW = Sjölunda Sewage Works (<LOD = below limit of detection)
A study carried out in the Rhine region of France, [Commission Internationale pour la
Protection du Rhin, 1999], showed that control of organic pollutants from point sources has
been effective at reducing levels of contamination in the Rhine. Between 1985 and 1996, the
pollution from PAHs and PCBs had decreased by over 90%. In 1985, 1,075 kg of PCBs were
discharged, which was reduced to 250 kg in 1992, and to 3 kg in 1996, all of which were
from industrial sources. For trichloromethane, 9,000 kg were discharged in 1985, 2,300 kg in
1992, and 2,210 kg in 1996; of these 600 kg were from industry and 1,610 kg from
communal sources.
68
3. Organic Pollutants
B.
DEHP
The Danish Ministry of the Environment and Energy [Danish Report, 1999] have estimated
the annual consumption of phthalates in Denmark to be approximately
o
o
10,000 tonnes in 1992 (about 90% of this used in soft PVC)
11,000 tonnes in 1995
In Germany, the total production of DEHP in 1988 was 234,000 tonnes. Of this, 1% was
discharged to surface and groundwater [Brüggermann, 1988].
The vast majority of phthalate emissions to the environment occur, not during the
manufacture, but during the use of the finished products. While in some cases this is a
commercial setting (such as vehicle washing, which will be examined subsequently), there
are also major sources in the domestic environment. Mattson et al. (1991) mentioned
previously regarding domestic sources of PAHs, estimated the household contribution of
phthalates and adipates to the Gothenburg sewage works as 70% of the total load (this
figure emphasises the ubiquity of compounds and difficulty of control). Two major sources of
domestic releases to wastewater (shown in bold in Table 3.3) are floor and wall coverings
and textiles with PVC prints.
Table 3.3 DEHP emissions in Denmark
[adapted from Appendix 1 Danish Ministry of Environment and Energy Report, 1999]
C
Product
Phthalate use (t y-1)
Emission to air
during production
(t y-1)
Emission to air
during use
(t y-1)
Release to
wastewater
during use
(t y-1)
Cars
Floor and Wall
Coverings
Textiles with
PVC prints
1000
2000
-
0.1-1
0.2
2-10
1-5
5-15
-
-
2-13
Dioxins and furans (PCDD/Fs)
The Environment Agency of England and Wales [1998] estimates dioxin emissions from
industrial Part A processes to UWW collecting systems in the UK as 4.5 µg (TEQ), whereas
emissions to air from these processes was estimated to be 1.1kg. Routes of these pollutants
into wastewater via deposition or industrial process (i.e. washing of air pollution cleaning
equipment), are not discussed. Actions taken to reduce dioxin emissions continue to ensure
IPC authorisations are met.
Recent research at the University of California, Berkeley, reports that deposition of dioxins to
soil is 6 to 70 times greater than estimated emissions [Eduljee 1999]. This suggests that
either not all sources of dioxin are known and/or the contributions from these sources may
not be accurately characterised.
Table 3.4 shows the dioxin emissions for the years 1994-1998 in Austria. There was little or
no change in the dioxin emissions in Austria over this period, but slight reductions, were
achieved in some sectors. The main reason for the emission reduction in 1998 is due to the
air hold ordinance, which limited dioxin emissions from waste combustion as well as from
steam-boiler plants.
69
3. Organic Pollutants
Table 3.4 Dioxin emissions in the time period 1994-1998, Austria
[Federal Environmental Agency, UNECE/CLRTAP, 1999].
1994
Issuer groups
Small consumer
(household, trade,
administration)
Industry (burning
and processes)
Industry
processes
Waste handling
and landfills
Total
1995
1996
1997
Dioxin emissions (tonnes per annum)
1998
16,820
18,160
18,400
16,780
16,260
3,470
3,730
3,880
3,980
3,910
8,170
8,900
7,990
8,550
8,380
180
180
180
180
180
28,640
30,970
30,450
29,500
28,740
In Spain, concentrations of dioxins are reported for recent samples (1999) of sewage sludge
and for archived samples (from 1979 to 1987) [Eljarrat, et.al., 1999]. Results are shown in
Table 3.5. It is estimated that the current concentrations of dioxin in sludge have dropped
since the 1970s-80s. This is expected to be due to the source reduction of pollutants, from
combustion and incineration processes, and from certain pesticides contamination and
emphasises the success that controls on use of compounds and trade effluent discharge in
reducing pollutant levels.
Table 3.5 Concentrations of PCDD/F in sewage sludge in Spain [Eljarrat, et.al., 1999]
Type of sewage
sludge
Fresh [1999]
Archived [1979-1987]
E.
Range of
concentrations
(pg.g-1 DW as I-TEQ)
7 to 160
29 to 8,300
Mean value
(pg.g-1 DW as I-TEQ)
55
620
Other organic compounds
Adsorbable organo-halogen compounds (AOX) resulting from bleach products and from
chlorine use, were reported in studies done in Portugal, in Ria Formosa lagooned sewage
[Bebianno, 1995] and in Italy in the city of Parma [Schowanek, et.al, 1996]. The average
AOX concentration in sewage was reported as 37 µg.l-1.
Sterols were reported in sewage sludge and around discharge wastewater points in
Portugal, in Faro, Tavira and Olhao [Mudge et al., 1997, 1998 and 1999]. Concentrations
ranged between 0.1 to 27.8 µg.g-1 sterols of dry weight of sludge. Hospital wastewater may
contain high phenol concentrations, up to 20,000 µg.l-1, plus other compounds such as LAS,
NPE, PCBs and pharmaceuticals.
F.
Vehicle washing
A specific activity identified as a source of a number of organic pollutants in urban
wastewater is vehicle washing, which consists of two distinct phases:
o
o
Actual cleaning, involving the removal of oily dirt, which, on a quantitative basis
would be expected to be similar to the type of oily dirt (asphalt and vehicle exhaust
particles) which is in road runoff. However, this would also involve the use of
degreasing solvents and surfactants which can enter the wastewater treatment
process.
Vehicle Treatment, involves the use of protective treatments, often coatings using
different types of wax against corrosion, dust and dirt.
70
3. Organic Pollutants
The effluent is usually discharged to the UWW collecting system. Several studies of the
effluents from vehicle washing facilities have been undertaken [Paxéus 1996a, 1996b,
Paxéus and Schröder, 1996, Ulmgren 2000a]. In Sweden, an environmental standard for car
washing detergents was established in Göteborg in 1992 [EHPA, 1992], based on the
Precautionary Principle and Substitution Principle in the Chemical Products Act. In general
COD values found at the effluents of vehicle washes are in the range of typical untreated
industrial petrochemical wastewaters [Huber, 1988].
In Gothenburg, an important site for vehicle manufacture, vehicle washing was estimated to
correspond to 0.5 % of the total wastewater at the Gothenburg WWTS, which was concluded
to have a very small effect on the total load of organic pollutants at the plant. The major
components of the effluents were aliphatic hydrocarbons and alkylbenzenes, originating from
petroleum base degreasing solvents and the oily dirt on the vehicles themselves (asphalt,
vehicle exhaust particles). Low aromatic products reduce the potential environmental
associated with detergent use in car washing facilities. These are produced by
hydrogenation of petroleum-based solvents where substituted benzenes and naphthalenes
are converted to corresponding naphthenes and decalins. The formation and discharge of
polyaromatic compounds is negligible for detergents that come from low aromatic
microemulsions.
Table 3.6 summarises the results of a study on washing both of light vehicles (LV) and
heavy vehicles (HV) [Paxéus 1996]. As can be seen, HVs tend to contribute larger organic
pollutant loads than LVs.
71
3. Organic Pollutants
Table 3.6 Concentration of organic pollutants in car wash effluents in mg l -1
[after Paxeus, 1996]
Conventional
parameters
LV
HV
Total oil
COD
Mean
291
1263
C8-C16
C17-C30
29
0.6
Benzene
Toluene
Naphthalene
0.01
0.08
0.17
Biphenyl
0.015
Dibenzofuran
0.001
Phenathrene
0.005
Pyrene
0.003
Fluoranthene
0.003
Diethyl phthalate
Dihexyl phthalate
0.005
0.05
DEHP
0.52
p-nonylphenol
2-Botoxyethanol
0.60
25
Median
242
1180
Range
10-1750
120-4200
Aliphatic hydrocarbons
22
1-139
0.4
<0.001
Aromatic hydrocarbons
0.01
<0.01-0.2
0.05
<0.01-0.6
0.13
<0.0010.7
0.005
<0.0010.1
0.002
<0.0010.03
<LOD
<0.0010.03
<LOD
<0.0010.01
<LOD
<0.0010.01
Plasticizers
0.01
2E-3-0.06
0.03
<0.0010.15
0.38
0.03 - 4.1
Washing agents
0.26
0.01-4
15
<0.001270
Mean
550
4600
Median
460
4500
Range
65-1200
17007500
103.86
1.84
76.72
1.87
41-220
0.9-3.0
0.02
0.10
1.1
0.02
0.08
0.75
0.02-0.03
0.03-0.2
0.3-3
0.12
0.11
0.04-0.2
0.011
0.011
0.021
<LOD
0.009
<LOD
0.004
<LOD
0.0090.012
0.0050.03
0.010 .02
0.0020.006
0.01
0.3
0.01
0.21
1.50
1.30
0.43
15
0.41
17
0.01-0.02
<0.0010.7
0.4 - 3
0.1-0.8
<0.00127
It is not known if this area is representative of the Scandinavian region as a whole in terms of
the car washing input. However, it does seem that car washing is also an important source
of pollutants in Norway [SFT, 1998a, 1998b]. In Norway 41 businesses were reported on as
sources of hazardous organic pollutants, PAHs, phthalates (DBP, BBP, DEHP),
nonylphenols (nonylphenol, nonylphenol mono- and di-ethoxylates). The studies found the
highest pollutant loads in the effluents from motor vehicle workshops to urban wastewater
came from petrol stations with car washes, long haul transport depots with ‘car washes’
commercial laundries, paint spraying workshop and chemical businesses [SFT, 1998a,
1998b].
There are two main types of washing agent available and the choice of these would result in
significant differences in wastewater quality:
• Water-based formulations (microemulsions) containing 10-30% hydrocarbons but
increased surfactants (10-30%);
• Petroleum-based degreasing formulations containing 95-99% of hydrocarbons and
3% surfactants.
Plasticisers found in the effluents from vehicle cleaning included phthalates, although
analysis of the cleaning and washing chemicals showed that they themselves contribute very
little to the discharge of plasticisers.
72
3. Organic Pollutants
3.1.2 Urban runoff
A significant proportion of organic contaminants in wastewater are derived from urban runoff.
These organic compounds include aliphatic and aromatic hydrocarbons, PAHs, fatty acids,
ketones, phthalate esters, plasticisers and other polar compounds. Solvent extractable
organics are dominated by petroleum hydrocarbons, which arise from motor oil and tyres
from road surfaces. Organic pollutant sources have not received the extent of research
attention that potentially toxic element pollution has. For example, in the case of PAHs which
are combustion by-products and enter wastewater principally through atmospheric
deposition and urban runoff, the sources can be stationary (industrial sources, power and
heat generation, residential heating, incineration and open fires) and mobile (petrol and
diesel engine automobile) [Sharma et al.,1994]. Different PAH species are associated with
each one of these sources.
A.
Road and vehicle related pollution
The main sources of road and vehicle related metals pollution have been outlined in Section
2.1.3. Table 3.7, shows some of the road and vehicle related sources of organic pollutants.
Table 3.7 Qualitative classification of road related sources of organic pollutants
[after Montague and Luker, 1994].
Traffic
Petrol
(PAHs and MTBE)
Oil
Grease
Antifreeze
Hydraulic fluid
Maintenance
Tar and bitumen
Accidents
Petrol
Oil
Grease
Solvents
PAHs
Asphalt
PCBs
Pesticides and
herbicides
Oil
Grease
Solvents
Table 3.8 summarises the results from three experimental catchments from 1975 to 1982 on
mean concentrations of PAH.
Table 3.8 Summary of pollutant concentrations in urban runoff caused by road related
sources [after Klein, 1982]
Pollutant mean
concentrations (mg.l-1)
Pleidelsheim
PAH
2.61
Test catchments
Obereisesheim
2.97
Ulm / West
2.51
The necessary conditions for PAH formation is the presence of benzene and a high
concentration of radical intermediates, which then form stable compounds. Multiple ring
systems are autocatalytic and promote further ring condensations. Fuel aromatic content has
been shown to influence particle-associated PAH emissions almost linearly [Pedersen et al.,
1980; Nunnermann, 1983; Egeback and Bertilsson, 1983]. However, the relationship
between the aromatic content of petrol and PAH formation is not fully understood.
PAHs are produced by unburned fuel, exhaust gases and vapour, lead compounds (from
petrol additives) and hydrocarbon losses from fuel, lubrication and hydraulic systems.
Volatile solids will be added to the total suspended solids loading of rainfall runoff and can
also act as carriers for both potentially toxic elements and hydrocarbons. Some road dusts
have been found to contain 8.5 µg g-1 of PAHs [Colwill et al., 1984 as reported in Luker and
73
3. Organic Pollutants
Montague, 1994]. The introduction of the catalyst technology for motor vehicles lowered the
emissions of PCDD/F in Germany to about 98% [UBA, 1999].
Tyre wear releases hydrocarbons either in particulate form or in larger pieces as a result of
tyre failure. A tyre loses about 10 to 20 per cent of its weight in a lifetime. Annually it is
estimated an average of 140 g of tyre-derived particles are eroded per metre of road
[Environment Agency of England and Wales, 1999].
Plasticisers (such as diethyl phthalate and dihexyl phthalate) are also considered an
important parameter of organic pollution load in urban runoff. Cary et al. [1989], stated that
plasticisers, especially phthalates, represent the major pollutants found in urban storm water.
The concentrations found for 8 plasticisers were recorded. Of these DEHP was found in the
greater concentrations than the other seven plasticisers combined. The main sources of
plasticisers are traffic grime and dirt, associated with the degradation of plastic components
of the vehicles.
B.
Roof Runoff
Regarding roof runoff as an interface between atmospheric boundary layer and the runoff
receiving system, Förster (1993) investigated the role of roofs as source and sink of organic
pollutants. The trace organics analysed included PAH, chlorinated hydrocarbons and nitro
phenols. The research indicated that the insecticide HCH was primarily introduced to the
roof runoff system by wet deposition, while the amount of adsorbed PAHs (pyrene;
benzo[a]pyrene=BaP) in roof runoff exceeded the input by rain with events during colder
times of the year where fossil fuel heating systems constitutes additional source for this
pollutant. The concentration profiles for a number of PAHs are illustrated in Figures 3.1 and
3.2 below.
Figure 3.1 PAH in runoff from zinc sheet roof [after Förster, 1993]
74
3. Organic Pollutants
Figure 3.2 PAH in runoff from tar roof [after Förster, 1993]
As can be seen, the concentrations of PAH in roof runoff from zinc roofs was found to be
about ten-fold higher than for tar roofs. There is a difference in the pattern of distribution for
PAH concentration at different precipitation flow rates. For tar roofs PAH concentration is
highest at the lower and higher precipitation flows and lower at intermediate events, whereas
for zinc roofs it tended to be higher at lower precipitation flows. Therefore, concentrations of
**pollutants in roof runoff can be considered variable depending on the characteristics of the
roof material itself as well as on the characteristics of the precipitation event.
A number of hydrocarbons are present in urban rainfall runoff, particularly those associated
with motor vehicles, such as petrol, fuel oils and lubricants. In an unmodified form these
liquids are insoluble in, and lighter than, water. Typically, 70-75% of hydrocarbon oils show a
strong attachment to suspended solids [Luker and Montague, 1994]. PAHs have an even
greater affinity. In contrast, Methyl-tertiary-butyl-ether (MTBEs) the new additive to unleaded
fuel is significantly more soluble in water than all other hydrocarbons in rainfall runoff.
Hydrocarbons, even in low concentrations, can give rise to surface sheens and thus
adversely affect surface waters. Most hydrocarbons eventually degrade by a combination of
microbial and oxidative processes; degradation though is slow, so the increase in oxygen
demand in watercourses and wastewater is likely to be marginal and not a principal
environmental impact.
C.
Urban vegetation control practices
Herbicides and pesticides are used in road maintenance operations to control weeds and
pests on the roadsides and verges. The triazine group of herbicides, including atrazine and
simazine, has been used extensively for roadside weed clearance and is more soluble and
mobile than their organo-chlorine predecessors. Combined levels of atrazine and simazine
above 1µg l-1 are not uncommon in watercourses near highways (Ellis, 1991). Collins and
Ridgeway (1980), report that half of pesticides in urban runoff are associated with particles
<63 µm, although these particles are less than 6% of the total suspended solids load.
In urban areas, pesticides in general, and herbicides in particular, are becoming an integral
part of the control of unwanted vegetation by local and municipal authorities, rail and airport
operators. The main herbicides used in the UK are of the triazine group, the phenyl urea
group (e.g. chlorotoluron, isoproturon and diuron), the phenoxy acid group (e.g. Mecoprop
and 2,4-D) and glyphosate (Revitt et al., 1999). Of the phenyl urea compounds, only diuron
75
3. Organic Pollutants
has been widely used in the urban environment and in 1989 this herbicide accounted for
13% of the total 550 tonnes of active ingredient used in the UK (Department of the
Environment, 1991). The comparable use of triazines was 39% but following the introduction
of restrictions for the non-agricultural use of these herbicides in 1992, many users converted
to the use of diuron and glyphosate for the control of vegetation in urban environments
(White and Pinkstone, 1995). The removal of herbicides by rainfall runoff is influenced by
rainfall characteristics, the time interval between herbicide application, the precipitation event
and the properties of the herbicide. However, the full range of factors that influence herbicide
release from sites of application and the mechanisms governing the transport to, and fate of
herbicides in the aquatic environment are not fully understood [Davies et al., 1995; Heather
and Carter, 1996]. The principal herbicide sources in urban catchments include [Revitt et al.,
1999]:
•
•
•
Urban parks and private gardens
Road maintenance (to road kerbstones and backwalls)
Railway system maintenance.
Concentrations in receiving waters, reported by Revitt et al., (1999) in the UK, were
consistently above the drinking water limit of 0.1 µg l-1 recommended for simazine and
diuron; the mean concentrations of which reached 0.34 and 0.45 µg l-1, respectively. In
France [Farrugia et al., 1999], the average application rates for pesticides on the most
consuming urban land uses are reported as 900 g ha-1 for roads and streets, 4000 g ha-1 for
cemeteries and 500 to 800 g ha-1 for parks and sport yards. Householders may also use
large amounts of herbicides and other pesticides but information on the quantities applied is
not available in published literature. However, there was considerable variation in the extent
of water contamination with herbicides between catchments. Farrugia et al, (1999), reported
the average concentration of diurons in water receiving urban runoff was 5 µg l-1, and
attributed this entirely to use in urban situations.
It is to be noted that the hydrological characteristics of hard urban surfaces provide the ideal
conditions for the efficient transport of herbicides (particularly diuron, see also Farrugia et al.,
1999) into UWW collecting systems. This, combined with the existence of inert physicochemical environments involving neutral pH, low nutrient and total organic carbon levels,
absence of absorption sites and low bacterial populations, allow the application of herbicides
in urban areas (although in low use), to be an important potential source of contamination of
waste water.
D.
Wet and dry deposition
The main repository of PCBs, PAHs and PCCD/Fs is soil. Volatilisation from soil, then further
atmospheric transport and deposition of PAHs, PCBs and PCDD/Fs is considered to be one
of the main contemporary sources of these contaminants in the environment Wild et al.,.
[1995]. PAHs are difficult to control because they are a combustion product.
The Austrian Federal Environment Agency (UBA) analysed PAHs in several media (surface
and wastewater, sediment, soil, sewage sludge, compost, plants, street dusts and ambient
air) between 1989 and 1998 [Gans, et.al., 1999]. Only 10 % of samples were above the
detection limit for PAHs of between 2.6 and 20.3 ng l -1 and these were all taken during winter
and spring, suggesting that PAH originates from the emissions of heating systems during the
cold period.
Once released (by the sources mentioned in the previous paragraphs), airborne PAHs are
transported by the prevailing meteorology before being removed from the atmosphere
through various scavenging mechanisms. As with other airborne pollutants the major
mechanisms of removal of PAHs from the atmosphere are wet deposition, such as rain,
sleet, snow, hail, and dry deposition to the surface. The wet removal of gaseous compounds
is better understood than particulate PAH removal [Ligocki et al., 1985]. The extent of in-
76
3. Organic Pollutants
cloud or below cloud scavenging, collection efficiency of falling precipitation, solubility and
size particles has been examined in the literature [McVeety, 1986 as reported in Sharma et
al., 1994].
Dry removal is a function of atmospheric conditions and the surface level concentration of
PAHs. PAHs adsorbed to particles greater than 20 µm have higher settling velocities and
thus will settle in the vicinity of the source. However, this mechanism will only account for a
minor percentage of removal, as PAH are mostly adsorbed on particles less than 10 µm in
diameter.
77
3. Organic Pollutants
3.2 INFLUENCE OF VARIOUS TREATMENT PROCESSES ON THE FATE OF ORGANIC
POLLUTANTS THROUGH WASTEWATER TREATMENT AND SEWAGE SLUDGE
TREATMENT
3.2.1 PARTITIONING OF ORGANIC POLLUTANTS IN WASTEWATER TREATMENT
PROCESSES.
The general effect of wastewater treatment processes is to concentrate the organic
pollutants in the sewage sludge and the extent of this removal depends on the properties of
the organic species. The overall result of this process is to discharge a treated wastewater
relatively free of organic and inorganic contaminants and a sewage sludge that contains
most of the organic contamination present in the feed wastewater. The main complication of
this general study arises from the large number of possible organic species that could be
present in the feed stream and the complex chemistry sorbtion mechanisms on the solids.
During the treatment cycle, some organic materials can degrade to a certain extent,
especially in aerobic environments and organic material of biological origin is easy to
degrade. Indeed, some common organic pollutants such as LAS, are specifically added to
detergents because they are aerobically biodegradable. A considerable body of literature
exists on this aspect and a variety of oxidants have been proposed. The main aim of this
type of work has concentrated on reducing the organic pollutant content in sewage sludge
prior to land disposal. Advanced oxidation processes might be used in tertiary treatment
especially if the final effluent is to be used for drinking water. However, use of these
processes; regardless of the power of the oxidant, cannot be expected, a priori, to degrade
all types of organic pollutants within a reasonably short time scale. Indeed, the presence of
organics in final effluents is an obstacle in expanding the recycling of wastewater.
3.2.2 Wastewater Treatment
Traditionally, wastewater treatment is supposed to begin at the head of a WWTS at the inlet
screens used to remove large objects such as wood plastics and paper. However, in reality
wastewater conditioning starts in the sewer, in large conurbations the wastewater can have
quite a significant residence time in a sewer. However, it is suggested that dilution of
sewage with runoff water is likely to have an adverse effect on the efficiencies of the
downstream treatment processes (Dorussen et al., 1997).
Primary treatment is installed to enable sedimentation of the feed wastewater. This process
is used to settle, retain and concentrate most of the particulate material to the bottom of the
tank as primary sludge. The process is affected by temperature and the solids content of the
supernatant or primary overflow is significantly higher if the temperature is low, as it is in
winter. Though simple, primary sedimentation is a widespread process in Europe, although
not practised in all WWTS. In some cases primary sedimentation is not installed and in other
plants flocculation, by addition of flocculants, is carried out in the primary sedimentation tank
(Hahn et al., 1999).
The objective of secondary treatment is to contact the primary overflow (settled sewage) with
air in the presence of aerobic bacteria and other micro-organisms, which convert the organic
matter to carbon dioxide and water to a variable extent. There are two types of plant
commonly used for this process: bio filters and activated sludge. Most WWTS use primary
and secondary processes. However some plants may have tertiary treatment which, can
involve coagulation, flocculation and rapid gravity filtration.
A novel process for secondary treatment is the lagoon (Salter et al., 1999). This is large unit
several meters deep and can be stirred gently and aerated. Aquatic life including fish can
survive in some lagoons. The residence time in the lagoon is long and they can be used to
treat the more contaminated municipal wastes. In addition secondary pre-treatment can be
carried out using magnetic flocs. In this process the organic contaminants present are
78
3. Organic Pollutants
loaded on to the magnetic flocs at a low pH and washed off in a high pH medium (Booker et
al., 1996).
There is some concern about the use of iron coagulants, which is of direct relevance to this
study. Some iron reagents used in wastewater treatment are made as a by-product of
titanium oxide production. The titanium ore contains traces of vanadium and uranium. Two
other tertiary methods often cited are activated carbon and membrane filtration. Both
however are rather expensive. Activated carbon is a very efficient means of removal of
organic pollutants and the technique is widely used in small domestic plants used to polish
drinking water. Membrane filtration is also very effective in removing particulate material
from water. However, the membranes are expensive and fouling can occur.
In order to estimate organic and inorganic pollutant removal in wastewater treatment
processes models are required to simulate them. In such models physical properties of the
pollutants are used to determine the likelihood that they will be removed by the process.
More work is needed on modelling the fate of organic pollutants through WWTS and their
transformation throughout the different treatment methods.
It is clear that the regular screening of priority organic pollutants on a day-to-day basis would
be complex and uneconomical. It has been suggested that determination of adsorbable
organic halogens (AOX) be used as an indicator for these priority substances (Hahn et al.,
1999). AOX determination is a relatively easy technique to use (Korner, 2000). These
substances are sorbed from the water on charcoal, which is subsequently pyrolysed. The
hydroxyhalides produced are sorbed and analysed by titration. Another general test
mentioned in the literature (Ono et al., 1996) is the bacterial umu-test, which measures
damage caused by organic pollutants on DNA.
3.2.3 Properties of Organic Pollutants
Octanol-water partition coefficient and solubility
The octanol- water partition coefficient is the ratio of a compound’s concentration in octanol
to that in water at equilibrium.
Concentration of compound in octanol
Kow =
Concentration of compound in water
Kow is dimensionless and values vary over the range of at least 10-3 to 107 and are usually
expressed logarithmically. Large Kow values are characteristic of large hydrophobic
molecules which tend to be associated with solid organic matter while smaller hydrophilic
molecules have low Kow values. Octanol-water partition coefficients can be measured directly
by using conventional “shake flask” methods (Leo and Hansch, 1971). This experimental
approach is restricted to compounds of low-to-medium hydrophobicity, since for compounds
with high hydrophobicity, the concentration in the aqueous phase is too low to be measured
accurately.
Kow can also be correlated with various environmental parameters, such as solubility. By
definition, the partition coefficient expresses the concentration ratio at equilibrium of an
organic chemical partitioned between an organic liquid and water. This partitioning is, in
essence, equivalent to partitioning the organic chemical between itself and water. One would
expect that a correlation would exist between the partition coefficient and solubility. Lyman et
al. (1990) presented the following correlation between solubility based on 156 compounds:
log
1
= 1.339 log K ow 0.978
Sw
where Sw is the solubility expressed in mol l-1. This correlation was obtained empirically and
the correlation coefficient was found to be 0.874.
79
3. Organic Pollutants
Organic carbon-water partition coefficient, KOC
The tendency of a compound to sorb to the organic matter such as humic substances
in soil or sewage sludge particles can be assessed using the organic carbon-water partition
coefficient. It is defined as the ratio between the concentration of the organic compound on
organic carbon (mg.g-1) and its concentration in water (mg.l-1), at equilibrium.
Concentration of compound on organic carbon
Koc =
Concentration of compound in water
The likelihood of the leaching of a compound through soil or adsorption onto soil organic
carbon can be assessed from K oc values. Generally, organic compounds with high K oc values
will tend to adsorb onto organic carbon whilst compounds with low values have a greater
tendency to be leached. Koc values can be estimated from the octanol-water coefficient or
water solubility. Karickhoff et al. (1981) found the following correlation:
Log Koc = 0.82 log Kow + 0.14
when he examined sorbtion data for a variety of aromatic hydrocarbons, chlorinated
hydrocarbons, chloro-S-triazines and phenyl ureas. The correlation coefficient was 0.93.
In this study a specific list of organic pollutants has been defined and it can be seen that
their solubilities are very low but the Koc values are very high in the region of 105 indicating
that the sorbtion would be very favourable. From the Koc values and the weight fraction of
organic carbon species present in the feed (f) an estimate of the removal of organic species
can be made. The amount left in the supernatant water as a percentage left (L) is given by:

1 

 K oc f 
L = 100 
Thus if f = 10-3 and Koc = 10-5, the percentage left would be 1%. There is limited data
available or actual results but figures for L are generally much higher. (Pham et al., 1997)
report that 30% of PCB and only 25% of the PAHs were removed from a specific treatment
plant.
3.2.4 Modelling
Understanding the processes involved in wastewater treatment is likely to provide a basis for
understanding the pathways and partitioning of pollutants in these processes. The way to do
this is to develop models of the processes and to simulate plants using computers. An
example of such a comprehensive model has been published (Gabaldon et al., 1998). The
model does not specifically include large molecular weight organics.
It is of some of interest to note that there is some work on processes that occur in a sewer.
One study aims to model the emissions of volatile organic compounds in cocurrent air flow in
open and closed sewers (Olsen et al., 1998). Another study measures the removal of COD
and proteins within a sewer (Raunkjaer et al., 1995) and found that there were quite
noticeable losses in a sewer. In another study the sewer pipe was considered to consist of a
sediment above which was a bio-film and above that the water phase (Fronteau et al., 1997).
O’Brien et al. [1995] and Mann et al.[1997] present a first order model for a wastewater
plant. In the secondary section aeration for stripping, biodegradation and sorption on to a
PAC (Powdered Activated Carbon) were considered. PCB, PCDD/F or PAH were not
included but the methodology presented in this paper could be applicable to the study of the
fate of these high molecular weight pollutants in secondary treatment. Work has been done
on modelling trickling filter-beds (Shandalor et al., 1997). This predicts the drop of solids
loading in the water as it trickles down the bed. On the more specific case of organic
80
3. Organic Pollutants
pollutant removal, a detailed paper has been published on the removal of volatile organic
contaminants in a wastewater plant (Melcer et al., 1994). However, again no specific
mention of PAH or similar organics was made.
3.2.5 Organic Degradation in Wastewater
Among the organic pollutants being studied in this report, LAS is somewhat unusual as it is
added to water in detergents. Studies in this subject (Holt et al., 1998 and Prats et al., 1997)
report very similar LAS degradation levels of over 90%. Although the removal of LAS in
WWTS is quite effective some 16% of the feed LAS is taken out in the sewage sludge (Field
et al., 1995). In this sorbed form it is more difficult to degrade. Some studies of LAS in river
sediments (Tabor et al., 1996) show that this compound is sorbed on to the solids and only
slowly biodegradable. Thus there would be an amount of non degraded LAS in the solid
residue.
There have been a number of studies on the degradation rate of PCDD, PCDF and PCBs,
which have been reported in a review article (Sinkkonen et al., 2000). The experiments were
conducted in laboratory rigs and the data reported as half-life analogous to radioactive
decay. The mean half-life quoted is given in Table 3.09.
Table 3.9: Half-lives of PCDD, PCDF and PCB in water
Substance
PCDD
PCDF
PCB
Half life in water
(years)
2.6
5.0
9.3
This study seems to indicate that these organics will not be degraded in a WWTS. These
half-lives are considerably longer than the residence time in a sewage treatment plant or
sewer. As the authors point out the experiments were conducted near ideal conditions and,
in practice, the half lives are believed to be longer than the figures quoted in the table,
especially if the temperature is low.
PAH compounds are believed to be persistent in the environment. There is some work that
presents evidence that some of these compounds can be degraded in periods of 12-80
hours (McNally et al., 1998). Compared with PCDDs this time period is rapid. However,
these experiments on biological degradation of PAH were carried out under ideal conditions.
There was a constant temperature (20oC), specially adapted bacteria were used and
nutrients were added. In a practical case where low temperature and few nutrients are
present, the actual degradation times would be much longer (in the region of 80-600 hours)
so PAH compounds are unlikely to be degraded in a conventional wastewater treatment
plant. Research in Greece by Samara et.al. [1995] and Manoli et al. [1999], shows that the
lower-molecular mass PAHs are removed effectively in Thessaloniki's WWTS, whereas the
higher molecular mass PAHs are resistant to the biological treatment. The heavy molecular
mass PAHs are partially removed by adsorption, whereas the lower molecular mass PAHs
are removed by volatilisation and/or biodegradation.
Work on oestrogenic compounds, analysing 17β-oestradiol equivalent concentrations, found
that the load of oestrogenic activity in the wastewater was reduced by about 90% in the
sewage plant. Less than 3% of the oestrogenic activities was found in the sludge (Korner et
al 2000).
81
3. Organic Pollutants
3.2.6 Removal of Organics
Coagulants such as aluminium and ferric salts are used in water treatment to remove
particulate matter. However, soluble organics may also be removed by coagulation by
mechanisms such as specific adsorption to floc particles and co-precipitation (Semmens and
Ocanas, 1977). Sridhan and Lee (1972) studied the removal of phenol, citric acid and
glycine from lake waters by co-precipitation with iron. Though these results were reasonable,
excessive concentrations of coagulant (300-1500 mg.l-1) were required. Other workers made
similar findings. Semmens and Ocanas (1977) examined the removal of dihyroxybenzoic
acid (DHBA) and resorcinol from distilled water by coagulation with ferric sulphate. Results
indicated that the extent of organic removal increased as coagulant dosage increased.
Maximum percentage removals were 35% for DHBA and 8% for resorcinol. Semmens and
Ayers (1985) examined the effectiveness of alum and ferric sulphate in removing octanoic
acid, salicylic acid, phenol and benzoic acid from Mississippi river water and water samples
free of organic matter. These compounds were generally poorly removed by coagulation and
in most cases the extent of removal did not depend strongly on coagulant dosage. Removals
ranged between 3-20%. Salicylic acid was most efficiently removed and benzoic acid was
most poorly removed. Generally, better removal of the organic compounds occurred when
natural organics were not present.
The general consensus of the work done to date indicates that the use of coagulants for
removing organics is feasible. However it is impractical as the excessive addition of
coagulants is necessary.
Humic substances account for around 50% of the dissolved organic matter in natural water
(Vik and Eikebrokk, 1989). They are formed easily from waste material and there is evidence
that they will sorb organic matter by binding with them. Activated carbon is widely used as a
means of removing organic compounds from water. The presence of humic acid reduces the
rate of organics uptake (Kilduff et al., 1988). The capacity of activated carbon for
trichloroethylene (Summers et al., 1989, Wilmanki and Breeman, 1990), trichlorophenol
(Najm et al., 1996) and lindane decreased in the presence of humic substances. Other
sorption media such as organoclays (Dentel et al., 1998, Zhoa and Vance, 1998) and an
organic polymer resin (Frimmel et al., 1999) are not so badly affected by the presence of
humic substances. Ying et al., (1988) studied the effects of iron precipitation on the removal
of natural organic compounds like tannic acid and humic acid, and toxic organic compounds
like chlorendic acid (HET), polychlorobiphenyls (PCBs) and organochlorine pesticides.
Freshly formed ferric hydroxide flocs were very effective in removing humic acid and tannic
acid and it was found that the presence of humic acid enhanced significantly the removals of
PCBs and many of the organochlorine pesticides by ferrous and ferric hydroxide
precipitates. Removals were achieved by a combined mechanism of complexation,
adsorption and co-precipitation. This evidence suggests that humic substances are capable
of sorbing organic material.
A process was devised in which organic contaminants were removed by adding humic acid
and a coagulant such as ferric hydroxide (Rebhun et al., 1998). This showed good recovery
for the organics tested. The results of this work suggest that humic acid might be added in a
tertiary cycle. The humic acid could be made by composting grass cuttings, potato peeling
and other waste feeds. Such material could be added to the final effluent of a wastewater
treatment plant followed by contact and flocculation.
3.2.7 Conclusions – removal of organics in wastewater
The practical issue of the removal of organics in wastewater treatment is not well
documented in the literature. Modelling work reviewed here, has shown that the work has
concentrated on the removal and degradation of organic matter of biological origin and that
synthetic organic pollutants have been largely neglected. Clearly modelling work for
pollutants should be promoted.
82
3. Organic Pollutants
This data in turn relies upon analysis of these organic pollutants. Present methods using
GC/MS are extremely complex and not suitable for routine plant use. Lack of easier methods
for their analysis will hinder the development of simple processes to remove these organic
materials. It could be argued that identification of a specific pollutant is not crucial for plant
development and that a generic test would be suitable. One of the most important aspects of
future work is the development or identification of simple tests for WWTP analysis. The
problem is not confined to treatment plants alone but rapid treatment methods could be used
to detect sources of heavy organic chemicals.
One of the results of the difficulty in doing analyses is that there is little data available on
partitioning process. There is a suggestion that around half the organics fed to a wastewater
treatment plant remain in the supernatant stream. This might well be a surprising result given
the measured property values in the region of 105 (dimensionless) would suggest that there
would be a good binding between organic pollutants and the settled sludge. It is possible
that there is some competition for sorbtion sites in the organic matter from the more
concentrated organic compounds.
With more rapid analysis techniques in place, there would be the opportunity to make
process changes to reduce the amounts of organics present in the final effluents. It is felt
that advanced oxidative techniques such as the use of ozone would not be applicable in the
present context as the organics have a very small concentration in solution and have a low
reactivity. One interesting possibility is in situ treatment in sewers such as adding activated
carbon to contaminated streams. As humic substances are efficient scavengers for organic
pollutants, humic acid derived from composting food waste could be added in a tertiary stage
to strip organics from the final effluent. It is a matter of policy, to see if such ideas should be
promoted further but initial work could start before the rapid analysis methods had been
agreed.
Transfer and partitioning of organic contaminants to the sludge matrix
The sorption of organic contaminants onto the sludge solids is determined by physicochemical processes and can be predicted for individual compounds by the octanol-water
partition coefficient (Kow). During primary sedimentation, hydrophobic contaminants may
partition onto settled primary sludge solids and compounds can be grouped according to
their sorption behaviour based on the Kow value as follows (Rogers, 1996):
Log Kow < 2.5
Log Kow > 2.5 and < 4.0
Log Kow > 4.0
low sorption potential
medium sorption potential
high sorption potential
Volatilisation and thermal degradation
Many sludge organics are lipophilic compounds that adsorb to the sludge matrix and this
mechanism limits the potential losses in the aqueous phase in the final effluent. A proportion
of the volatile organics in raw sludge including: benzene, toluene and the dichlorobenzenes
may be lost by volatilisation during wastewater and sludge treatment at thickening,
particularly if the sludge is aerated or agitated, and by dewatering. Volatilisation is used to
describe the passive loss of organic compounds to the atmosphere from the surface of open
tanks such as clarifiers. The majority of volatilisation, however, occurs through air stripping in
aerated process vessels. As a general guide, compounds with a Henry’s Law constant >10-3
atm (mol -1 m -3) can be removed by volatilisation (Petrasek et al., 1983). The significance of
volatilisation losses of specific organic compounds during sewage treatment can be
predicted based on Henry’s constant (Hc) and Kow (Rogers, 1996):
Hc > 1 x 10-4 and Hc/Kow > 1 x 10-9
Hc < 1 x 10-4 and Hc/Kow < 1 x 10-9
high volatilisation potential
low volatilisation potential
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3. Organic Pollutants
However, more recent studies (Melcer et al., 1992) suggest that the stripping of volatiles
may not be as significant as was initially thought and biodegradation during secondary
biological wastewater treatment may be the main mechanism of loss of the potentially
volatile compound types (Table 3.10). For example, Melcer et al. (1992) reported that
biodegradation processes removed ≥90 % of the dichoromethane, 1,1,1-trichoromethane,
trichloroethylene, toluene and xylene from a municipal wastewater. Volatilisation was only a
significant mechanism of removal for 1,4-dichlorobenzene (20 %) and tetrachloroethylene
(60 %). The fate and behaviour of volatile organic compounds in wastewater treatment plant
have been modelled numerically by the TOXCHEM computer-based model that incorporates
four removal mechanisms including: volatilisation, stripping, biodegradation and sorption on
to solids (Melcer et al., 1992).
Table 3.10 Observed and predicted (TOXCHEM) removals of volatile organic
contaminants during wastewater treatment by stripping and biodegradation (Melcer et
al., 1992)
Compound
Air stripping (%)
Biotransformation (%)
Observed
Predicted
Observed
Predicted
Dichloromethane
Chloroform
1,1,1-Trichloroethane
Trichloroethylene
Tetrachloroethylene
1,4-Dichlorobenzene
Toluene
p- and m-Xylene
2.6
7.4
10.5
10.7
58.7
19.1
1.2
1.3
3.2
7.8
6.0
3.1
64.2
17.2
0.4
0.6
92.4
73.6
79.7
82.7
15.8
54.7
98.6
98.1
91.9
71.9
89.1
91.3
0.0
54.8
98.3
97.9
High temperature treatment of sludge by disinfection processes at 70 oC for 30 minutes can
enhance the loss of volatile compounds. Mono- and two-ringed aromatic compounds
(benzene, toluene, xylene, naphthalene, dichlorobenzene etc) may be partially lost under
these conditions (Wild and Jones, 1989). Other more persistent hydrophobic compounds, eg
lesser chlorinated PCBs, and the three-ringed PAHs, may also be susceptible to
volatilisation. Thermal drying is being introduced as an enhanced treatment process to
produce sanitised biosolids for unrestricted use and for improved handling and bulk
reduction. This process is potentially the most effective at removing volatile substances from
sludge because the solids are exposed to high temperatures (400 oC) and the sludge is
dried to >90 % ds. Thermal degradation may also be an important mechanism for the
removal of organic contaminants from sewage sludge during heat treatment (Wild and
Jones, 1989). Volatile organic compounds in sewage sludge are not regarded as a potential
risk to human health or the environment when sludge is used in agriculture (Wilson et al.,
1994).
Destruction by sludge stabilisation processes
Mesophilic anaerobic digestion is the principal sludge stabilisation process adopted in most
European countries, where approximately 50 % of sludge production is treated by this
methods. Volatile compounds are generally lost to the atmosphere or transferred to the
supernatant during digestion, whereas PAHs and phthalate acid esters are conserved (Bridle
and Webber, 1982).
Many organic contaminants are biodegraded under anaerobic conditions and this is
enhanced by increasing retention time and digestion temperature. Five characteristic
behaviour patterns (Figure 3.3) of decay are observed for organic contaminants in anaerobic
digestion systems based on net gas (total CH4 + CO2) production (Battersby and Wilson,
1989):
• Easily degradable (eg ethylene glycol, diethylene glycol, triethylene glycol, sodium
stearate, ethanol);
84
3. Organic Pollutants
•
Degradable after a lag phase (eg phenol, 2-aminophenol, 3- and 4-cresol, catechol,
sodium benzoate, 3 and 4-aminobenzoic acid, 3-chlorobenzoic acid, phthalic acid,
dimethyl phthalate, di-n-butyl phthalate, pyridine and quinoline);
• No degradation or gas production (3- and 4-aminophenol, 2-chlorophenol, 2-cresol,
2-nitrophenol, 2- and 4-chlorobenzoic acid, bis (2-ethylhexyl)phthalate, hexylene
glycol, neopentyl glycol, n-undecane, n-hexadecane, 2,4-D, dieldrin, cis- and transpermethrin, tetrahydrofuran, furan, pyrrole, N-methylpyrrole, thiophene, benzene,
pyrimidine, 1-naphthoic acid);
• Inhibitory in the initial phase of incubation (eg 3- and 4-chlorophenol, 2,4- and 2,6dichlorophenol, 2,4,6-trichlorophenol, 3- and 4- nitrophenol, 2-phenylphenol, 2-, 3and 4-nitrobenzoic acid, CTAB, MCPA, MCPP, lindane, naphthalene,
anthraquinone);
• Inhibitory throughout incubation (eg 3,5-dichlorophenol, pentachlorophenol, 2,4- and
2,5-dinitrophenol, 4-nonylphenol, sodium dodecylbenzene sulfonate, sodium 4octylbenzene sulphonate, 2,4,5-T, butyltin trichloride, dibutyltin dichloride, tributyltin
chloride).
Degradation is generally aided by carboxyl and hydroxyl groups, whereas chloro or nitro
groups tend to inhibit anaerobic biodegradation and gas production.
Figure 3.3 Typical patterns of net gas production (CH4 + CO2) from organic chemicals
incubated anaerobically with diluted primary digested sewage sludge.
1, Easily degradable; 2, Degradable after a lag period; 3, little effect on gas production; 4,
inhibitory in initial phase of incubation; 5, inhibitory throughout incubation
(Battersby and Wilson, 1989).
1
2
}3
Time
Inhibition
Net gas production (% theoretical)
100
-100
4
5
Biodegradation during anaerobic digestion may virtually eliminate certain organic
contaminants from sewage sludge, but in general the destruction achieved is typically in the
range of 15 – 35 % (WRc, 1994). Aromatic surfactants including linear alkyl benzene
sulphonates (LAS) and 4-nonylphenol polyethoxylates (NPnEO) occur in sludge in large
concentrations. These compounds are not fully degraded during sewage treatment and there
is significant accumulation in digested sludge. For example, mass balance calculations
suggest that approximately 80 % of LAS is biodegraded during the activated sludge process
and 15-20 % is transferred to the raw sludge (Brunner et al., 1988). Approximately 20 % of
the LAS in raw sludge may be destroyed by mesophilic anaerobic digestion sludge. The
compounds, nonylphenol monoethoxylate (NP1EO) and nonylphenol diethoxylate (NP2EO)
are formed during sewage treatment from the microbial degradation of NPnEO. These
85
3. Organic Pollutants
metabolites are relatively lipophilic and accumulate in the sludge and are also discharged
with the treated sewage effluent. One of the most important consequences of anaerobic
digestion, however, is the production of nonylphenol (NP), which accumulates in digested
sludge. Approximately 50 % of the NPnEO in raw sewage is transformed to NP in digested
sewage sludge. The loadings of LAS and NP to soil in sewage sludge used on farmland are
significantly larger than for most of the other organic contaminants present in sludge and
there is concern about their potential environmental effects. This is particularly the case for
NP in sludge due to its potential oestrogenic activity (UKWIR, 1997). However, in the aerobic
soil environment, these compounds provide substrates for microbial activity and are rapidly
degraded so there is minimal potential risk to the environment or transfer to the human
foodchain. For example, LAS has a short half-life in soil in the range 7 – 22 days in
temperate field conditions (Holt et al., 1989) and the half-life for NP is <10 days (UKWIR,
1997). Current studies at Imperial College, funded by the Food Standards Agency in the UK,
are investigating the potential for plant uptake of NP into staple food crops from sludgetreated soil.
Another class of organic chemicals, the phthalate acid esters, are also an abundant group of
compounds present in sewage because of their extensive use as plasticising agents. The
phthalates are also suspected as being potential environmental oestrogens (UKWIR, 1997).
Shelton et al. (1984) reported the complete degradation of the lower molecular weight
phthalate esters, and of butyl benzyl phthalate, within 7 days in laboratory scale anaerobic
digesters operated at 35 oC. Therefore, these phthalate compounds should generally be
removed by most municipal anaerobic digesters at the normal mean retention times
operated in practice (>12 days). The extent and rate of biodegradation during anaerobic
digestion is apparently related to the size of the alkyl side chain and compounds with larger
C-8 group are much more resistant to microbial attack. Therefore, di-n-octyl and di-(2ethylhexyl)phthalate (DEHP) are considerably more persistent to anaerobic microbial
mineralisation and are generally not removed by conventional anaerobic stabilisation
processes. However, phthalate esters are rapidly destroyed under aerobic conditions,
usually achieving >90% removal in 24 h in activated sludge wastewater treatment systems.
In soil, the reported half-life is <50 d (UKWIR, 1997).
Composting is a thermophilic aerobic stabilisation process and usually involves blending
dewatered sludge at approximately 25 % ds with a bulking agent, such as straw or wood
chips, to increase porosity of the sludge to facilitate microbial activitiy. The biodegradation of
relatively persistent organic compounds such has been reported for composted sludge (Wild
and Jones, 1989). For example, PAHs may be partially degraded by composting sludge and
average removals of 13 % and 50 % have been measured for benzo(a)pyrene and
anthracene, respectively, although phenanthrene persisted unchanged in laboratory
composting trials (Martens, 1982; Racke and Frink, 1989).
Thermophilic aerobic digestion processes and sludge storage for three months can achieve
similar overall removal rates for organic contaminants as those obtained with mesophilic
anaerobic digestion (WRc, 1994). Thermal hydrolysis conditioning of sludge prior to
conventional anaerobic stabilisation may have a significant influence on the removal of
organic contaminants from sludge, but this is a comparatively new enhanced treatment
process and effects on the destruction of organic contaminants have yet to be investigated.
86
3. Organic Pollutants
3.3
Quantitative assessment of organic pollutants in untreated UWW, treated UWW
and treated SS
For the main list of organic pollutants considered in this report there is little available data of
the concentrations in the influent to the wastewater treatment plant. Paxéus and Schröder
[1996] looked at over 50 organic compounds, in the influents and effluents of the
Gothenburg wastewater treatment plant. The high cost of testing explains the lack of data on
dioxins in urban wastewater.
Most of these compounds were reduced to below the limit of detection during the treatment
process. Some of the organic compounds, such as caffeine were reduced from a level of
37µg.l-1 to 4µg.l-1. Some of the phosphorus containing compounds were not reduced during
the treatment process (although the influents and effluents were quite low at 1µg.l-1). The
overall toxicity of the influent and the effluent were also measured and found to have
decreased by approximately 50% during the treatment process.
Dioxin concentration (ng TEQ
kg-1 ds)
450
400
German limit =
100 ng TEQ kg-1 ds
350
300
250
200
150
100
50
0
1944
1949
1953
1956
Year
1958
1960
1998
Figure 3.4 Dioxin content of archived samples of sewage sludge form Mogden WWTS,
UK
It can be seen (Figure 3.4) that there has been a significant reduction in the concentration of
dioxins since the 1950s and 1960s in sludge over recent years.
The concentrations of other organic contaminants in sludge, including, PCBs and PAHs,
have also declined significantly in sludge in the UK. This is due to the control of primary
sources of these substances. In 1984, McIntyre and Lester (1984) measured median and
99th percentile concentrations for PCBs in sludge (444 samples from UK sewage treatment
works) of 0.14 and 2.5 mg kg-1, respectively. Ten years later, Alcock and Jones (1993)
reported the total PCB content of 12 UK sludges from rural, urban and industrial sewage
treatment works ranged between 0.106 to 0.712 mg kg-1, with a mean value of 0.292 mg kg1
. These results indicate that overall PCB concentrations in UK sludges have declined
markedly in response to the ban on industrial production, use and discharge of these
substances. Similar trends are apparent in Germany (Table 3.11). In effect, this means that
the chemical composition of sewage sludge is already subject to stringent, albeit indirect,
controls that have been effective in minimising industrial sources and inputs of persistent
organic contaminants.
87
3. Organic Pollutants
Table 3.11 Mean concentrations of organic contaminants in German sewage sludge in
1988/89 relative to data collected until 1996 (Leschber, 1997)
Contaminant
Adsorbable organo-halogens mg kg-1 ds
Polychlorinated biphenyls(1) mg kg-1 ds
Polycyclic aromatic hydrocarbons(1) mg kg-1
ds
Di(2-ethylhexyl)phthalate mg kg-1 ds
Nonylphenol mg kg-1 ds
Dioxins and furans (ng TEQ kg-1 ds)
(1)
1988/89
250-350
<0.1
0.25-0.75
1991/96
140-280
0.01-0.04
0.1-0.6
50-130
60-120
<50
20-60
15-45
Single congeners
Table 3.12 Survey of organic pollutants in UWW and WWTS (µg.l-1)
Compound
PAHs
Country
Austria:
Total PAHs - EPA15
Germany:
Total PAHs
Benzo(a)pyrene
Benzo(k)fluoranthene
Greece:
Benzo(a)pyrene
Fluoroanthene
Indeno (1,2,3-cd) pyrene
France
Germany
Anionic
Surfactants
Detergents
LAS
NPE
Reference
Influent
(µg.l-1)
Effluent
(µg.l-1)
147-625
20-70
0.79
0.08
0.05
Gans et al.,1999
Hagenmaier et al,
1986
Manoli et al, 1999
0.022
0.24
0.015
0.05-0.44
33
0.005
0.029
0.005
0.02-0.09
Austria
51.8 (5.6 to
349)
4.4
30.8 (2.4147)
0.3
Germany
Italy
122 (7-232)
290-4800
15 (5.6-184)
-
France
Austria
Germany
Greece
1-26
400-3500
5400
Italy
Netherlands
Spain
UK
Austria:
Nonylphenolmonoethoxylate
Nonylphenoldiethoxylate
Italy:
NP
NPEO
NPEC
Sweden
Germany
4600
4000
9600
15100
0.1-2.7
11-55
67
129 (35325)
43
9
140
10
2,096,000
363,000
13,093,000
639,000
UK
DEHP
WWTS
4
27
145
0.5-6.0
0.02
0.002
ADEME, 1995
Koch et al, 1989 &
Balzer et al 1991
Morris et al, 1994
Hohenblum et al.,
2000
Faltin, 1985
Braguglia et al,
2000
ADEME, 1995
Scharf et al., 1995
Feijtel et al 1995
Kilikidis et al. 1994
Feijtel et al 1995
Feijtel et al 1995
Feijtel et al 1995
Feijtel et al 1995
Hohenblum et al.,
2000
Di Corcia et al.
1994
Paxéus 1996a
Koppe et al, 1993
88
3. Organic Pollutants
Other organic
pollutants:
Chlorophenols
Chlorinated
organics
Pesticides
VOCs
France
0.1-0.4
300
1.15
10
Germany
<0.1-0.5
ADEME, 1995
Ternes et al, 2000
Iodinated X-Ray
contrast
substances:
iopamidol
diatrizoate
iothalamic acid
iomeprol
iopromide
Total Phenols
4.3
3.3
0.18
0.17
1.6
7.5
Italy
2.5-300
Dioxins
Italy:
0.024-16.9
Italian Regional
Environmental
Protection Agency
Italian Regional
Environmental
Protection Agency
PAHs: wastewater from 8 different sewage treatment influents was investigated in 1996 by
the UBA [Gans, et.al., 1999]. Similar PAHs content were determined, except for the influent
from a chemical plant, which had an approximately 1,000 times larger concentration. PAHs
especially, with a low molecular weight were found in high concentrations. Apart from the
higher PAH content of the wastewater from the chemical plant, no significant differences
could be detected between municipal and industrial influents. The PAHs content of the
effluent was about 10 times smaller than the influents [Gans et.al.1999].
The Danish regulation of the application of waste products [Ministry of the Environment and
Energy 1996] sets certain cut off values for the maximum concentrations of organic
contaminants in sludge to be distributed on agricultural land as shown in Table 3.13.
Concentrations of PAHs in Danish sewage sludge are also shown in Table 3.14 The PAH
concentration of the nine selected compounds were all found to have mean concentrations
above the concentrations permissible for use on agricultural land in Denmark.
Table 3.15 Danish standards for maximum concentrations of organic contaminants in
sewage sludge (Danish Ministry of the Environment and Energy, 1996)
Danish Standards
LAS
nonylphenol (including nonylphenol
ethoxylates)
PAHs*
DEHP
1997 - cut off values 2000 - cut off values
mg.kg-1 DS
mg.kg-1 DS
2,600
1,300
50
10
6
100
3
50
*(total concentration of nine selected PAHs) Acenaphthylene, Fluorene, Phenanthrene, Fluoranthene,
Pyrene, Benzo(b,j,k)fluoranthene, Benzo(a)pyrene, Benzo(g,h,i)perylene and Indeno(1,2,3,-cd)pyrene
The values in bold are difficult to achieve, as they are far below current sludge
concentrations. If 50% of pyrene and phenathrene is from food sources and gives sludge
concentrations of > 300mg.kg -1 ds, then this emphasises how difficult these standards are to
achieve.
The mean concentrations of LAS, NPE and DEHP were found to be within the Danish limits
for use on agricultural land but the range of concentrations in all cases went over the cut off
limits; therefore many of the sludges would not be allowed to be used on agricultural land.
89
3. Organic Pollutants
The concentrations of some of the organic contaminants in the sludge were found to depend
strongly on the wastewater treatment process [Danish EPA]. The concentrations of LAS, NP
and NPE were significantly lower (P<0.005) following activated sludge treatment than in
mixed activated and digested sludge treatment, presumably due to extended aeration.
It can be seen that government and other institutions are trying to introduce limits for certain
pollutants and that concern for wastewater pollution reduction is increasing. Nevertheless, it
is noted that important discrepancies exist in analysis techniques, even within a country,
hence slowing the determination of limits, particularly for PAHs and PCBs. Due to the
expected increase in sludge production and the reinforcing of the legislation in relation to the
concentration limits for potentially toxic elements and organic pollutants, it seems necessary
throughout Europe to harmonise analysis techniques and the pollutants targeted in the
control of wastewater and sludge quality. Discharge standards to UWW collecting systems
for industries and possibly reformulation of certain domestic products should be determined
in order to reduce pollution entry into the systems.
Table 3.14a) Survey of organic pollutants in sewage sludge: mg kg-1 DS (a)PAHs
PAHs
Country
Mean
Median
Min.
Max.
B[A]P
I[1,2,3-cd]p
Σ PAHs*
B[a]p
Fl.thene
I[1,2,3-cd]p
B[a]p
B[a]p
Fl.thene
I[1,2,3-cd]p
Fl.thene
B[a]p
Fl.thene
B[a]p
Fl.thene
I[1,2,3-cd]p
Σ PAHs
Fl.thene
Austria
0.30
0.27
6.4
0.35
1.2
0.3
0.5
0.15
0.3
0.67
3.4
0.22
0.21
0.24
1.1
0.11
1.2-2.2
0.24
1.3
0.12
0.7-1.4
0.09
0.07
2.6
0.1
0.6
0.1
0.1
<0.01
<0.01
<0.01
1.1
0.04
0.15
0.1
0.38
0.05
0.67
0.58
15.3
1.1
2.7
0.8
3.4
1.4
3.3
0.63
6.0
11
31
0.36
1.4
0.15
0.01
0.7
6.0
0.1
1.1
0.04
83.8
7.5
4
1.83
Germany
(municipal)
Denmark
Spain
France
Greece
Sweden
Σ PAHs**
B[a]p Fl.thene
UK
Σ PAHs***
Switzerland
(municipal)
Italy
B[a]p
I[1,2,3-cd]p
Fl.thene
Σ PAHs
B[a]p
Fl.thene
Σ PAHs****
Pyrene
B[a]p
Poland
EU
USA
Limits
EU
WHO
USEPA
0.4
0.07
0.1
0.23
27.8
2.6
2.5
0.35
2.3
0.50
<0.05
<0.05
<0.05
72.3
Year/s of
Survey
1994/95
(24)
1996 (10)
1996 (10)
1996 (10)
(10)
1995 (14)
(28)
(19)
1994 (6)
1994 (6)
(15)
1995/98
(26)
(29)
1994 (27)
1989 (33)
1991 (29)
(29)
2000 (34)
74.4
13.8
9.95
32.7
4.7
Agricultural Soils
114.3
1999 (2)
154
154
1988
(30/31)
Sewage Sludge
6 (proposed)
480
21.4
(9)
(5)
(25)
B[a]p: Benzo[a]pyrene;
Fl.thene: Fluoranthene
I[1,2,3-cd]p: indeno(1,2,3-c,d)pyrene
* sum of 16 USEPA priority list PAHs (see Appendix B)
** sum of naphthalene, acenaphthylene, acenaphthene, fluorene, phenanthrene, anthracene, fluoranthracene,
pyrene, chrysene, benzo(a)anthracene
*** sum of benzo(b+k)fluoranthene, benzo(ghi)perylene, benzo(a)pyrene, fluoranthene, indeno(1,2,3-c,d)pyrene
90
3. Organic Pollutants
**** sum of acenapthene, phenapthene, fluorine, fluoranthene, pyrene, benzo(b+j+k)fluoranthene,
benzo(a)pyrene, benzo(ghi)perylene, indeno(1,2,3-c,d)pyrene
Table 3.14b) Survey of organic pollutants in sewage sludge: mg kg-1 DS (b)PCBs
PCBs
Country
Mean
Median
Min.
Max.
PCB (28,52,101,
138, 153,180)
Austria
0.07
0.05
0.02
0.27
Germany
0.01-0.04
0.5
0.05
15
<0.03
0.05
0.03
0.2
0.93
0.4
Year/s of
Survey
1994/95
(24)
1991-96
(13)
1985-87 (7)
(28)
(20)
1994 (6)
21.5
1995/98
(26)
(18)
PCB
(101,118,138)
Denmark
Spain
France
0.05
0.03
Sweden
0.1
0.1
UK
EU
USA
Limits
EU
WHO
USEPA
0.34
1.46
0.01
1.48
14.8
Agricultural Soils
Sewage Sludge
0.8 (proposed)
30
6.6
1988
(30/31)
(9)
(5)
(25)
Table 3.14c) Survey of organic pollutants in sewage sludge: mg kg-1 DS (c)DEHP
DEHP
Bis-(2ethylhexyl)phthalate
Bis-(2ethylhexyl)phthalate
Bis-(2ethylhexyl)phthalate
Country
Mean
Austria
Germany
20-60
Denmark
Sweden
38
EU
Canada
USA
Limits
EU
110
Median
Min.
Max.
Year/s of
Survey
23.4
34.4
<2.4
320
25
3.9
6.7
170
28
(11)
1991-96
(13)
(7)
(28)
(29)
68.0
11
959
(3)
891
1988
(30/31)
17
Agricultural Soils
Sewage Sludge
100 (proposed)
(9)
91
3. Organic Pollutants
Table 3.14e) Survey of organic pollutants in sewage sludge: mg kg-1 DS (e)LAS
LAS
Aerobic
Anaerobic
Aerobic Anaerobic
Country
Mean
Median
Min.
Max.
Austria
Germany
Denmark
Spain
8107
5000
2700
7579
2199
50
11
100
12100
17955
16000
16100
500
17800
Finland
Italy
UK
EU
USA
9700
11500
60
14000
18800
Limits
8700
530
10400
152
4680
Agricultural
Soils
Year/s of
Survey
1994/95(24)
1985-87(7)
(28)
(1)
(21)
(17)
(4)
(12)
(16)
(22)
1680
7000
Sewage Sludge
EU
2600 (proposed)
(9)
Table 3.14f) Survey of organic pollutants in sewage sludge: mg kg-1 DS (f)NPE
NPE
Country
Mean
Median
Austria
Germany
24
60-120
51
20
10
15
12
NP1EO
NP2EO
Denmark
Sweden
13-27
326-638
UK
EU
USA
Limits
EU
Min.
Max.
69
8
400
10-26
Agricultural Soils
3.8
5
<3
0.3
26
96.3
80
80
67
1100
256
824
Sewage Sludge
50 (proposed)
Year/s of
Survey
1994/95(24)
1988/89(13)
1996 (10)
(7)
(7)
(28)
1990 (32)
1995/98(26)
(27)
(9)
Table 3.14g) Survey of organic pollutants in sewage sludge: mg kg-1 DS (g)PCDD/F
DIOXINS &
FURANS (NG
TEQ/KG DS)
Country
Mean
Germany
15-45
Spain
55
620
24
40.2
42
7
29
23
7.6
160
8300
25
192
1991-96
(13)
1994-98 (8)
1979-87 (8)
(23)
(29)
82.7
90.4
37.4
0.49
2321
1820
1988
(30/31)
Sweden
UK
EU
USA
Dioxins
Limits
EU
Median
Agricultural Soils
Min.
Max.
Year/s of
Survey
Sewage Sludge
100 (proposed)
(9)
92
3. Organic Pollutants
References
1.
2.
Berna JL et al, 1989
Bodzek, B. et al,
1999.
3. Bridle, T.R. et al 1983
4. Cavelli L, et al 1993
5. Chang, A.G. et al
1995.
6. Conseil supérieur
d'hygiène publique de
France, 1998
7. Drescher-Kaden et al
1992,
8. Eljarrat. E, et al 1999.
9. European Union, 2000
10. Hessische
Landesanstalt fur
Umwelt (1991-96).
11. Hohenblum, P, et al
2000.
12. Holt MS et al 1992.
13. Leschber, R. 1997
14. Litz. N, et al, 1998.
15. Manoli, E. et al 1999.
16. McAvoy DC, et al
1994.
17. McEvoy & Giger 1986
18. McIntyre. A,E, et al
1984
19. Moreda, JM, et al
(1998a)
20. Moreda, JM, et al
(1998b)
21. Prats D, et al 1993.
22. Rapaport RA, et al
1990.
23. Rappe et al 1989
24. Scharf, S, et al. 1997
25. Smith, S.R. 2000
26. Statistika
meddelanden 1998
27. Sweetman 1994
28. Tørsløv J, et al 1997.
29. UKWIR 1995
30. USEPA 1992
31. USEPA 1999
32. Wahlberg,.C, et al
1990
33. Wild. S,R, et al 1989.
34. Braguglia et al 2000
Table 3.15 shows the occurrence of certain organic pollutants in sewage sludge in Germany.
Table 3.15 Occurrence of certain organic substances in sewage sludge, Germany
[Priority list USEPA and 6/464/EEC of EG].
Compound
Occurrence
in sludge
Benzo(a)anthracene
+++
Benzo(a)pyrene
+++
Benzo(k)fluoranthene
+++
Dibenzo(a,h)anthracene
+++
Indeno(1,2,3-cd)pyrene
+++
PCB-1242
++
PCB-1254
+++
PCB-1221
++
PCB-1232
++
PCB-1248
++
PCB-1260
+++
PCB-1016
+
2,3,7,8-Tetrachlordibenzo-p-dioxin
++
Frequency of occurrence: +++ frequent (90-100%), ++ less frequent, + low frequency
93