ART the Biodegradation of Surfactants in the Environment

Biochimica et Biophysica Acta 1508 (2000) 235^251
www.elsevier.com/locate/bba

Review

The biodegradation of surfactants in the environment
Matthew J. Scott, Malcolm N. Jones *
School of Biological Sciences, University of Manchester, Oxford Road, Manchester M13 9PT, UK
Received 9 February 2000; received in revised form 22 June 2000; accepted 3 August 2000

Abstract
The possible contamination of the environment by surfactants arising from the widespread use of detergent formulations
has been reviewed. Two of the major surfactants in current use are the linear alkylbenzene sulphonates (LAS) and the alkyl
phenol ethoxylates (APE). These pass into the sewage treatment plants where they are partially aerobically degraded and
partially adsorbed to sewage sludge that is applied to land. The biodegradation of these and a range of other surfactants both
in wastewater treatment plants and after discharge into natural waters and application to land resulting in sewage sludge
amended soils has been considered. Although the application of sewage sludge to soil can result in surfactant levels generally
in a range 0 to 3 mg kg31 , in the aerobic soil environment a surfactant can undergo further degradation so that the risk to the
biota in soil is very small, with margins of safety that are often at least 100. In the case of APE, while the surfactants
themselves show little toxicity their breakdown products, principally nonyl and octyl phenols adsorb readily to suspended
solids and are known to exhibit oestrogen-like properties, possibly linked to a decreasing male sperm count and carcinogenic
effects. While there is little serious risk to the environment from commonly used anionic surfactants, cationic surfactants are
known to be much more toxic and at present there is a lack of data on the degradation of cationics and their fate in the
environment. ß 2000 Elsevier Science B.V. All rights reserved.
Keywords: Surfactant; Linear alkylbenzene sulphonate (LAS); Alkyl phenol ethoxylate (APE); Wastewater treatment; Sludge amended
soil; Environment

1. Introduction
Abbreviations: AS, fatty alcohol sulphates; AE, fatty alcohol
ethoxylates; AES, alcohol ether sulphates; APE, alkyl phenol
ethoxylates; DDT, dichloro diphenyl trichloroethane; DM, dialkyldimethylammonium chloride; DTDMAC, ditallowdimethylammonium chloride; FES, fatty acid esters; OECD, Organisation
for Economic Cooperation and Development; LAB, linear alkylbenzene; LAS, linear alkylbenzene sulphonates; NP, nonyl phenol ; NPE, nonyl phenol ethoxylate; PEC, predicted environmental concentration; PNEC, predicted no e¡ect concentration; PT,
propylene tetramer; SDS, sodium dodecyl sulphate; SAS, secondary alkane sulphonates; TM, alkyltrimethylammonium chloride; WWTP, wastewater treatment plant
* Corresponding author. E-mail: [email protected]

Detergents are formulations designed to have
cleaning/solubilisation properties. These formulations consist of surface-active agents (surfactants) together with subsidiary components including builders
(e.g. tripolyphosphate), boosters, ¢llers and auxiliary
compounds. In terms of environmental issues the focus of concern has largely been on the e¡ects of the
surfactant in a detergent formulation, although there
was a period when the increasing use of builders
presented problems. From 1947 to 1970 the use of
tripolyphosphates in the USA rose from approximately 100U103 tons pa to 100U106 tons pa before

0005-2736 / 00 / $ ^ see front matter ß 2000 Elsevier Science B.V. All rights reserved.
PII: S 0 3 0 4 - 4 1 5 7 ( 0 0 ) 0 0 0 1 3 - 7

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M.J. Scott, M.N. Jones / Biochimica et Biophysica Acta 1508 (2000) 235^251

Fig. 1. Surfactant consumption in the USA, Japan and western Europe during 1982 [94].

the introduction of restrictive legislation [1]. In this
review we will focus on the major surfactants which
pass into the environment and the potential problems
which can arise as a consequence. The bulk of the
materials reaching the environment (soil and natural
waters) do so from consumer products via the use of
sewage sludge on land, e¥uents from wastewater
treatment plants (WWTP) and industrial discharges
into freshwater and marine sites. Fig. 1 outlines the
major uses of surfactant in the USA, Japan and
western Europe. Other sources of surfactant contamination are the use of surfactant dispersants for fuel
oil spillages and surfactant-enhanced remediation of
subsoil after spillage and contamination with nonaqueous liquids.
Historically, potential surfactant contamination of
the environment followed the shift from the use of
soap-based detergents to synthetic surfactants. The
transition period was approximately the 30 years
from 1940 to 1970 when the use of synthetics rose
from 4.5U103 tons pa in the USA to approximately
4.5 a 106 tons pa, while the use of soap fell from
1.4U106 tons pa to 0.6U106 tons pa [1]. During
this time there was also a transition from the use
of solid domestic detergents (powders) to liquids.
Until 1960 the major surfactant used in detergency
was propylene tetramer benzene sulphonate (PT benzene). It was about this time when sewage treatment
problems began to arise and foaming problems arose
on rivers. PT benzene was being discharged into
water systems and was found to be resistant to biodegradation by bacteria due to the branched alkyl
chain. The prohibition of this non-biodegradable sur-

factant forced the switch to more biodegradable
straight chain alkyl surfactants and now the major
anionic surfactant in use is linear alkylbenzene sulphonate (LAS). The production of surfactants of
various types in the USA, Japan and western Europe
is shown in Fig. 2. The place of LAS in the detergent
industry has been reviewed by Schoenkaes [2]. In
1994 the production of LAS in the USA, western
Europe and Japan was 840U103 (metric) tons pa
[3] although consumption is currently falling as new
alternatives are introduced. LAS represent more than
40% of all surfactants used. It is not surprising that a
large part of the literature is focussed on the environmental problems arising from LAS.
Another widely used class of surfactant is the alkyl
phenol ethoxylates (APE), which are used in detergents, paints, pesticides, textile and petroleum recovery chemicals, metal working and personal products.
Worldwide production of APE is 500U103 tons pa
[4]. Commercial formulations usually contain mixtures of APE (di¡erent chains length and isomers)
but with high proportions of nonyl and octyl alkyl
groups. Restrictions on the use of APE have arisen
since the discovery in 1984 that their breakdown
products are more toxic to aquatic organisms than
the APE themselves. Biodegradation of APE leads to
the shortening of the ethoxylate chains to alkyl phenol carboxylates leading ultimately to nonyl and octyl phenols, which have low water solubility and adsorb to suspended solids and sediments. Nonyl
phenol (NP) is approximately 10 times more toxic
than its ethoxylate precursor [4]. It is known to
mimic the e¡ect of the hormone oestrogen. Nonyl

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237

Fig. 2. Production of the di¡erent types of surfactants used in the USA, Japan and western Europe in 1982 [94].

phenolics in wastewater extracted from digested sewage sludge can pass into rivers. It has been hypothesised that endocrine disruptors may be responsible
for a decreasing male sperm count, testicular and
breast cancers [5]. Sublethal toxic e¡ects of NP on
zooplankton in natural waters interferes with their
sex determination and development [6]. These problems are leading to bans and restrictions on the use
of APE for household and industrial cleaning applications in Europe but not in the USA. In the US
researchers are less convinced of the adverse e¡ects
of alkyl phenols, possibly because of di¡erences in
wastewater treatments in the USA which result in
higher removal rates as compared to Europe [4].
As the above indicates, the major problems of surfactants in the environment arise from the two major
classes of materials LAS and APE but in the case of
the APE the environmental problems relate more to
the biodegradation products rather than to the APE
themselves. In this review we will concentrate largely
on the LAS where e¡ects relate directly to the surfactant.
2. Biodegradation and ecological impact of surfactants
Balson and Felix [80] described biodegradation as
the destruction of a chemical by the metabolic activity of microorganisms. When reviewing the literature
concerning the degradation of surfactants it is apparent that studies quote ¢gures for primary and/or ul-

timate biodegradation. Primary degradation can be
de¢ned as to have occurred when the structure has
changed su¤ciently for a molecule to lose its surfactant properties. Ultimate degradation is said to have
occurred when a surfactant molecule has been rendered to CO2 , CH4 , water, mineral salts and biomass.
LAS are generally regarded as biodegradable surfactants. Very high levels of biodegradation (97^
99%) have been found in some WWTP using aerobic
processes [7^9]. In contrast, APE are less biodegradable and values of 0^20% have been quoted [10]
based on oxygen uptake and 0^9% based on spectroscopic techniques [10].
The mechanism of breakdown of LAS involves the
degradation of the straight alkyl chain, the sulphonate group and ¢nally the benzene ring [11,12]. The
breakdown of the alkyl chain starts with the oxidation of the terminal methyl group (g-oxidation)
through the alcohol, aldehyde to the carboxylic
acid as follows (see Fig. 3). The reactions are enzyme
catalysed by alkane monooxygenase and two dehydrogenases. The carboxylic acid can then undergo Loxidation and the two carbon fragment enters the
tricarboxylic acid cycle as acetylCo-A. It is at this
stage that problems arise with branched alkyl chains,
a side chain methyl group or a gem-dimethylbranched chain cannot undergo L-oxidation by microorganisms and must be degraded by loss of one
carbon atom at a time (K-oxidation).
The second stage in LAS breakdown is the loss of

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Fig. 3. The reaction pathways of g- and L-oxidation of the alkyl chain during surfactant degradation [80].

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Fig. 4. Transport of surfactants through a WWTP. The ¢gures
will vary from one plant to another. The ¢gures shown are approximate values taken from reported values (data from Dentel
et al. [17]).

the sulphonate group, although there has been some
discussion about the sequence of steps [11]. Three
mechanisms have been proposed for desulphonation
according to the following reactions.
Hydroxyative desulphonation:
RSO3 H ‡ H2 O ! ROH ‡ 2H‡ ‡ SO23
3

…1†

Monooxygenase catalysis under acid conditions:
RSO3 H ‡ O2 ‡ 2NADH ! ROH ‡ H2 O‡
‡
SO23
3 ‡ 2NAD

…2†

Reductive desulphonation:
RSO3 H ‡ NADH ‡ H‡ ! RH ‡ NAD‡ ‡ H2 SO3
…3†
Whichever mechanism prevails the breakdown
product of the LAS is sulphite which can be oxidised
to sulphate in the environment.
The loss of the alkyl and the sulphonate group
from LAS leaves either phenylacetic or benzoic acids.
Microbial oxidation of phenylacetic acid can result in
fumaric and acetoacetic acids and benzene can be
converted to catechol [11].
Studies on the biodegradation of LAS and other
surfactants by bio¢lms of bacterial populations isolated from riverine [13] and estuarine [14] sites have

239

been reported. A study of biodegradation of a range
of anionic surfactants at a river site (river Ely, South
Wales, UK) located near a sewage treatment plant
outfall has been made [13]. Experiments were conducted in the laboratory using a population of bacteria isolated from river stone bio¢lms. Water collected at the out£ow (BO), upstream (BU) and
downstream (BD) of the site was incubated with
the isolated bacteria. It was found that the reciprocal
half-life (or `die-away' time) for biodegradation
of surfactants followed the sequence alkyl sulphates
s alkyl ethoxy sulphates s secondary linear alkyl sulphates s primary alkane sulphonates s LAS, and
that `die-away' time of surfactants depended on the
site in the sequence BO s BD s BU [13]. The ability
of bacterial species in the population to biodegrade
sulphonated surfactants was less widely distributed
than the ability to biodegrade sulphate ester surfactants.
In a study of LAS biodegradation by bacterial
cultures originating from an estuarine site (Krka river estuary, Croatian Mid Adriatic region; a highly
strati¢ed karstic estuary) it was found that the rate of
biodegradation depended on the origin of the culture, temperature and the structure of the alkylbenzene group [14]. Cultures isolated from the freshwater layer of the river had a greater ability to
degrade LAS than those from the underlying saline
water layer. Degradation rates were faster for the
longest alkyl chain LAS (in this study C13 ), and
slower for LAS isomers having the sulphophenyl
group situated in the middle of the alkyl chain.
The complete biodegradation of surfactants requires
a consortium of bacteria due to the limited metabolic
capacities of individual microorganisms [15]. The opportunity for commensalism (bene¢t to one microorganism) and synergism to develop exists in a consortium. Such interactive e¡ects lead to more
e¡ective biodegradation than is possible by any individual microorganism. The biodegradation of LAS
requires a four membered consortium, three members of which oxidise the alkyl chain but synergism
amongst the four members was essential for mineralisation of the aromatic ring [16].
A large amount of surfactant is associated with
sewage sludge solids. However, LAS are not biodegraded by either mesophilic or thermophilic anaerobic digestion. Various estimates of the load of LAS

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M.J. Scott, M.N. Jones / Biochimica et Biophysica Acta 1508 (2000) 235^251

and APE in a typical WWTP and their subsequent
fate have been given by Dentel et al. [17]. Fig. 4 is
based on data quoted by Dentel et al. [17]. Although
there are wide ranges in some of the values, e.g.
values quoted for the load of LAS in treatment
plants range from 3 to 21 mg l31 , it is clear that
signi¢cant amounts of surfactant are transported
into the environment from treatment plants. Dentel
et al. [17] estimate a discharge of over 100 kg day31
of anionic surfactants and approximately 300 kg
day31 of cationic surfactants from a 90U106 gal
day31 WWTP. It should also be noted that the presence of surfactants in water at concentrations below
and above the critical micelle concentration can also
lead to the solubilisation of other oil-soluble pollutants such as DDT and trichlorobenzene [18]. The
problem of the analysis of surfactants in the aqueous
environment has recently been reviewed by Lukaszewski [19]. With the introduction of new types of
surfactants developed to replace ethoxylates such as
alkyl polyglucosides it is necessary to have methods
for their detection and for the detection of their
breakdown products as well as improving existing
methods for speci¢c determination of di¡erent
classes of anionic and non-ionic surfactants. The behaviour of LAS in sewage by using direct UV absorption spectra deconvolution has been described by
Djellal et al. [20].
The assessment of risk to the environment from
surfactant contamination and surfactant catabolites
is an important and by no means a simple issue [21^
23]. A study in the Netherlands [21] of the risk to the
aquatic environment from a range of surfactants and
soaps, placed the materials in the priority order LAS,
alcohol ethoxylates, alcohol ethoxylated sulphates
and soap. The study looked at the ratio of the parameters `predicted environmental concentration'
(PEC) at 1000 m below the sewage outfall to the
`predicted no e¡ect concentration' (PNEC). The
data were gathered from seven locations and supplied by Netherlands industries. For the surfactants
the PEC/PNEC ratio was 0.05 but for soap it was
almost 1. These results suggest that little risk to the
aquatic environment is expected.
The risk assessment of LAS to terrestrial plants
and animals reported by Mieure et al. [23] also concludes that there are adequate margins of safety in
the use of wastewater for the irrigation of plant spe-

cies. The most vulnerable plant species are orchids
and vegetables grown hydroponically (radish, Chinese cabbage and rice). Adverse e¡ects on plant
and animal species (earthworms Eisena foetids and
Lumbricus terrestris) were observed at LAS concentration of 10 mg l31 , however LAS concentrations in
wastewater e¥uents are in a range 0.09 mg l31 to 0.9
mg l31 . These ¢gures give a safety margin in a range
10 to 100. The e¡ect of surfactant on plant growth
from the use of sewage sludge is di¤cult to assess
because in general the sludge promotes plant growth.
Adverse e¡ects on plant growth were observed at 392
Wg g31 but long term monitoring at a range of 46
environmental sites gave LAS concentrations of 6 3
Wg g31 . These ¢gures give a safety margin of 131.
For terrestrial animals the limit of no adverse e¡ects
was 235 Wg g31 giving a safety margin of 78.
However, in looking at ecotoxicity from sewage
e¥uents the less toxic surfactant residues and surfactant catabolites must be considered and this requires
analytical tests for these entities [22]. The monitoring
of LAS and their degradation products in the marine
environment, especially from littoral zones, is more
complex due to potential interference from other natural surfactants and other organic compounds [24].
Some of the problems of surfactant detection and
estimation in the environment may be solved in the
future by the use of speci¢c biosensors. An optical
biosensor for the determination of ionic surfactants
based on the immobilisation of acrylodan labelled
bovine serum albumin onto silanised silica optical
¢bre has been developed. It had a linear dynamic
range from 5 to 60 WM and a response time of less
than 30 s, although the long term stability of the
biosensor needs improving [25].
The biodegradation of LAS is e¡ected by a number of factors amongst which are the concentration
of dissolved oxygen [26], complexing with cationic
surfactants [27,28], the formation of insoluble calcium and magnesium salts [29], the presence of other
organic contaminants [30,31] and the e¡ect of LAS
on the pH during aerobic degradation [32]. In sewage-contaminated groundwater the rates of LAS biodegradation increase with dissolved oxygen concentration and the longer alkyl chain homologues (C12
and C13 ) are preferentially biodegraded. However,
the removal of LAS was found to be 2^3 times greater under laboratory conditions than in ¢eld tracer

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studies [26]. The formation of LAS complexes with
cationic surfactants (alkyltrimethylammonium chloride (TM) and dialkyldimethylammonium chloride
(DM)) leads to complex adsorption onto river sediments [27]. The adsorption of complexes which form
with molar ratios of LAS to cationic surfactant in
the range 1:1 to 6:1 (with TM) and 1:1 to 2:1 (with
DM) obey the Freundlich adsorption isotherm. The
rates of biodegradation as measured over 14 days for
1:1 and 2:1 complexes relative to the rate of biodegradation of LAS (taken as 100%) were as follows:
2LAS:TM (56%), LAS:TM (36%), 2LAS:DM (31%)
and LAS:DM (29%). The kinetics of biodegradation
of LAS and other organic matter by mixed bacterial
cultures as used in activated sludge treatment can be
a¡ected by LAS at high concentrations ( s 20 mg
l31 ). This arises as a consequence of LAS decreasing
the pH during aerobic degradation [32]. The highest
rates of biodegradation are found for the longest
alkyl chain homologues.
The biodegradation of APE by bacteria in seawater polluted with urban sewage is brought about
by bacteria of the Pseudomonas genus of marine
origin. Few other species of Gram-negative bacteria
are able to degrade APE with nine^ten ethoxy
groups. Pseudomonas strains degrade only down to
four or ¢ve ethoxy groups, although other species of
bacteria which are unable to degrade the long chain
APE are able to degrade the APE with four or ¢ve
ethoxy groups down to the two ethoxy group compounds [33].
3. Surfactants in sewage sludge
The literature concerning the fate of surfactants in
sewage sludge amended soil is heavily biased towards
the study of LAS with other surfactants receiving
little or no attention.
Due to their amphiphilic nature surfactants in raw
sewage can adsorb to the surface of resident particulate matter. Surfactants may also precipitate from
solution in the presence of metal ions (particularly
Ca2‡ ). Such behaviour may result in a signi¢cant
proportion of the surfactant load of raw sewage
being associated with the particulate fraction. A
common initial step in a WWTP is the removal of
particulate matter in primary settling tanks. Sludge

241

collected from these tanks is relatively rich in surfactant. Treatment of such sludge is commonly anaerobic digestion at elevated temperature. Many common
surfactants used are easily biodegradable in aerobic
conditions but due to restricted metabolic pathways
the majority are not degradable under anaerobic
conditions. Therefore, sludge treated anaerobically
may still be relatively rich in surfactants post treatment. Matthews [34] reported that during 1977 of the
1.3 M tons of sewage sludge produced in the UK
45% was disposed of as a fertiliser to agricultural
land. The remaining 55% was disposed of via land¢ll
sites or incineration. Anaerobically digested sludge
(dry weight) can contain 0.3^1.2% LAS [35^37].
The addition of anaerobically digested sewage sludge
to agricultural land is a large potential source of
LAS and other surfactants to the soil environment.
3.1. Linear alkylbenzene sulphonates
Berna et al. [38] reported that a signi¢cant proportion of LAS in raw sewage (10^35%) adsorbs to particulate matter. Sediment removed from primary settling tanks is relatively rich in LAS, with
concentrations ranging from 5000^15 000 mg l31
being reported [7,35,37,39,40].
The process of adsorption of LAS to particulate
matter is primarily driven by the hydrophobic e¡ect
and speci¢c or electrostatic interactions [41]. The extent of adsorption has been shown to be dependent
upon a number of factors. Prats et al. [42] suggested
that the type of LAS homologue present might be
signi¢cant. Longer alkyl chains conferred greater hydrophobicity thus increasing adsorptive tendency.
Painter [43] stated that for each carbon atom added
to the alkyl chain a two- to three-fold increase in the
Ka (association constant) for LAS was observed.
The characteristics of the water carrying the e¥uent can have a signi¢cant e¡ect upon the adsorption
of LAS. Berna et al. [38] showed that water hardness
could signi¢cantly alter partition coe¤cients of LAS
in raw sewage. Waters high in Ca2‡ concentrations
yielded sludge from primary settling tanks that contained 30^35% of the LAS concentration of the raw
sewage, but relatively soft water yielded only 10^
20%.
The presence of high concentrations of LAS in
sewage sludge leaving the WWTP is dependent

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M.J. Scott, M.N. Jones / Biochimica et Biophysica Acta 1508 (2000) 235^251

Table 1
Fate and persistence of surfactants in sludge amended soils
Application
form

Location

Surfactant/
derivative

Soil
concentration
post
application
(mg kg31 )

Monitoring
period

Final soil
concentration
(mg kg31 )

Half-life
(days)

Author

Sludge onto soil

Spain

LAS

22.4

Prats et al. [42]

Switzerland

LAS

45

3.1
0.7
5

not reported

Sludge onto soil

6 months
12 months
12 months

9

Marcomini
et al. [53]

Surfactant onto
soil

Germany

NP
LAS

4.7
not reported

2 months

0.5
not reported

5^25 summer

Litz et al. [92]

Sludge onto soil

Germany

LAS

16

6 months
76 days

0.19

68^117 winter
13

Surfactant onto
soil

USA

LAS

27
0.05

106 days
40 days

0.44
not reported

26
1.1^3.6

Sludge onto soil

Spain

LAE
LAS

Sludge onto soil

UK

LAS

90 days
170 days
5^6 months

0.3
not reported
61

26
33
7^22

Sludge onto soil

UK

LAB

55 days

0^0.38

15

Holt and
Bernstein [45]

Composted wool
scour sludge

Australia

NPE

14 weeks

1200

not reported

Jones and
Westmoreland [81]

0.05
16
53
2.6^66.4
(estimated
cumulative
load)
0.3^9.5
(estimated
cumulative
load)
14 000

upon the type of treatment the sludge undergoes. It
is well reported that LAS is readily degradable under
aerobic conditions. The alkyl chain oxidation at the
terminal methyl group (g-oxidation) requires the
presence of molecular oxygen. Subsequent degradation of the chain (L-oxidation) is followed by oxidative ¢ssion of the aromatic ring to give sulphonatesubstituted dicarboxylic acids. Finally, desulphonation of ring degradation products occurs [23,44^
46,89]. The g-oxidation of the alkyl chain and the
cleavage of the benzene ring require molecular oxygen, therefore under anaerobic conditions degradation via these pathways is unlikely. At present no
evidence exists for comparable degradation of LAS
under anaerobic conditions [47^49]. Holt et al. [44]
stated that sewage sludge is generally digested under
anaerobic conditions. Jensen [50] compiled results

Figge and
Scho«berl [93]
Knaebel et al. [86]

Berna et al. [52]
Waters et al. [51],
Holt et al. [44]

from ten studies of LAS in treated sewage sludge
from various locations around the world. He found
that sewage sludge that had been aerobically treated
had LAS concentrations of 100^500 mg kg31 dry
weight. This was considerably lower than levels
found in anaerobically treated sludge (5000^15 000
mg kg31 dry weight). Therefore, the extent of LAS
contamination of sewage sludge is greatly dependent
upon the individual WWTP and the method of
sludge digestion it employs.
During the past 15 years the fate of LAS in sludge
amended soils has received a great deal of attention.
Table 1 shows a summary of the results reported
from studies examining the fate and persistence of
surfactants. The most comprehensive studies were
carried out by Holt et al. [44], Holt and Bernstein
[45] and Waters et al. [51]. These studies examined

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LAS and LAB (un-reacted precursor of LAS constituting approximately 1^3% of LAS concentration) in
agricultural sludge amended soils. Comparisons were
made between estimated cumulative soil load of
LAS/B (calculated from previous sludge application
amounts and LAS/B contents of those sludges) and
soil LAS/B concentration. These studies encompassed 51 ¢elds on 24 farms situated in the Thames
Water Authority region of southern England. The
sample sites included di¡erent soil types (clay, silt
clay, clay loam and sandy loam) and agricultural
use (arable or pasture); sludge of di¡erent compositions and origins (homologue distributions; 2.7^34%
dry solids; 15^341 mg LAS l31 sludge); frequency of
application (0^6 annually), post application soil concentrations (0^293 mg LAS kg31 soil) and methods
of sludge application (subsurface injection or surface
dressing). Results showed that the 42 ¢elds that had
not been treated within 1987 (the year of the studies)
exhibited soil concentrations of LAS of 0^2.5 mg
kg31 (83% containing 6 1 mg kg31 ). For the majority of sites this constituted a s 98% removal of LAS
when compared to the cumulative estimated load.
Nine ¢elds that had been treated within 1987 showed
soil concentrations of 0.2^19.8 mg kg31 compared to
estimated cumulative loads of 15^206 mg kg31 , losses
of LAS constituted 70^90% of the estimated cumulative load. Five ¢elds that were monitored in a time
course to determine disappearance rates of LAS
showed degradation half-lives to be approximately
7^22 days. Holt et al. [44] concluded that degradation was primarily microbially driven and that soil
type, agricultural land use, application method and
whether a soil had been ploughed or not had no
e¡ect upon degradation rates. LAS homologue distribution showed no signi¢cant changes post application suggesting no di¡erential degradation. LAB was
found to have a soil residence half-life of approximately 15 days. Fields with applications during 1989
(year of sampling) exhibited concentrations of 5^390
Wg kg31 soil and ¢elds not receiving an application
during 1989 generally had concentrations 6 5 Wg
kg31 (constituting 99% loss).
Berna et al. [52] monitored LAS concentrations in
a sludge amended Spanish grapevine farm and a vegetable farm. From relatively high sludge application
concentrations of 7000^30 200 mg kg31 dry weight
initial soil concentrations of 16 and 53 mg kg31

243

soil respectively were observed. After a period of
90 and 170 days the soil concentration of LAS was
0.3 mg kg31 , calculated half-lives were 26 and 33
days.
Marcomini et al. [53] observed that after an initial
period of LAS removal soil concentrations appeared
to level out and not decrease further. Soil levels post
sludge application rapidly fell from 45 mg kg31 soil
to 5 mg kg31 , from which point no further signi¢cant
changes were observed. The authors suggested that
this observation might be due to the LAS being incorporated into the soil particles and/or being associated with the soil organic matter. This e¡ectively
rendered them unavailable to the microorganisms responsible for their biodegradation.
From the studies carried out it is evident that once
applied to the aerobic soil environment LAS is readily degradable with a half-life of 1^87 days. Once
removed from the anaerobic environment of sludge
digestion and/or storage, bacteria begin to metabolise
LAS. Rapid metabolism leads to relatively short
half-lives of LAS. Most authors who have carried
out monitoring of LAS residence in sludge amended
soils agree that due to their relatively high biodegradability in the aerobic environment there exists
little chance of accumulation of LAS in soil.
3.2. Soap
Soap is still a commonly used surfactant. Steber
and Berger [54] reported that C12 ^C18 soaps are readily available to microbes. However, the poor solubility of soaps precipitated with metal ions in£uences
biodegradation rates. Soaps are susceptible to precipitation within hard water environments. Scho«berl et
al. [55] reported that Sturm-tested sodium soap salts
exhibited mineralisation rates of 80^90% whereas calcium soap salts were signi¢cantly less biodegradable
with only 67% mineralised. Association of soaps with
metal ions leads to precipitation of a signi¢cant
quantity of soap. Precipitation leads to sedimenting
of soap in the primary settling tanks at WWTP.
Therefore, depending upon the hardness of water at
an individual WWTP sewage sludge may contain a
signi¢cant amount of soap. Anaerobic degradation
of soap is an important factor in determining the
e¡ectiveness of WWTP and the degradation of these
compounds. The main degradation pathway involved

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in soap breakdown is the L-oxidation of the alkyl
chain [54]. This step does not require the presence
of molecular oxygen and therefore anaerobic conditions should not hinder soap degradation. Steber and
Weirich [56] placed a14 C16 soap in an anaerobic bioreactor for 28 days and observed 14 CO2 and 14 CH4
production. Results suggested that the soap was 92^
97% degraded during that time period. Birch et al.
[48] observed a 79^94% degradation of sodium palmitate during a 3^4 week period of anaerobic digestion of sludge. The data suggest that soap is readily
biodegradable in both aerobic and anaerobic environments and therefore is ultimately treatable within
the conditions and residence time of wastewater
within a WWTP.
3.3. Secondary alkane sulphonates
When compared to LAS, SAS have had few direct
measurements made upon their biodegradability. Little is known about their degradation pathways and
their resistance to biodegradation in the WWTP. In
the aerobic environment SAS are considered to be
readily biodegradable. Swisher [10], Scho«berl et al.
[55] and Painter [43] have all reported rapid primary
degradation of SAS with s 90% removal in less than
3 days. Lo«tzsch et al. [58] incubated uniformly labelled C17 SAS and observed 61% ultimate degradation, producing CO2 . Neufahrt et al. [59] incubated
14
C17 SAS for 3 days observing 47% degradation to
CO2 and 25% incorporation into the biomass. Both
studies suggest that SAS are readily ultimately degradable under aerobic conditions. Steber and Berger
[54] assuming that SAS had similar degradation
characteristics to LAS stated that SAS might not
be available during anaerobic digestion. Bruce et al.
[57] observed that SAS concentrations did not appear to fall during anaerobic digestion of sludge. It
is suggested that as LAS and SAS have similar molecular characteristics their degradation pathways
may be similar. Therefore, a lack of molecular oxygen will inhibit primary g-oxidation of the alkyl
chains and oxidative desulphonation. Like LAS,
SAS is readily degradable in aerobic conditions but
a proportion may adsorb to particle surfaces in the
raw sewage and be removed in primary settling
tanks. If such sludge is not aerobically digested
then SAS concentrations in sludge may well be rela-

tively high. Like LAS it is thought that once
amended to a soil SAS in sludge will be rapidly biodegraded in the aerobic conditions, suggesting
chance of accumulation in soil is remote. However,
until monitoring studies of SAS in sludge amended
soils are performed such statements have to be acknowledged as conjecture.
3.4. Fatty acid esters
Fatty acid esters (FES) are readily degradable
under aerobic conditions. Gode et al. [60] observed
a 99% primary degradation of FES in an OECD
(Organisation for Economic Cooperation and Development) screening test, and a 76% ultimate degradation in a closed bottle test. Similar results were reported by Steber and Weirich [61] who outlined
possible degradation mechanisms including g-oxidation of the terminal alkyl group followed by L-oxidation of the alkyl chain and a subsequent desulphonation of the resultant short chained carboxy ester
sulphonate. Steber and Berger [54] point out that this
is a similar degradation mechanism as that for LAS
and therefore FES may be poorly degradable under
anaerobic conditions. Maurer et al. [62] observed
FES concentrations in an anaerobic digester, commenting that no primary degradation was observed
within the 30 days experimental period. Steber and
Weirich [61] observed that of a 14 C labelled FES
digested anaerobically for 4 weeks less than 5% of
the label was observed as CO2 or CH4 production.
This suggests that any FES associated with sludge
particles may pass through a WWTP being relatively
untreated. Steber and Weirich [61] did report that
FES in anaerobically treated sewage sludge was
quickly degraded when applied to aerobic soils.
3.5. Fatty alcohol sulphates
Fatty alcohol sulphates (AS) are rapidly aerobically degraded. Swisher [57] reported primary degradation ¢gures of 100% after 1+ day(s). Steber et al.
[63] reported primary degradation rates of 95^98% in
a 5 days period. Scho«berl et al. [55] reported primary
degradation ¢gures of 99% for an OECD test and
ultimate degradation ¢gures of 64^96% in a closed
bottle test. Swisher [57] stated that as biodegradation
of AS is so rapid it could be assumed that the pro-

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cesses involved in the primary and ultimate degradation are e¡ective in a broad range of microorganisms. The degradation of AS is thought to involve
the enzymatic cleavage of the sulphate ester bonds to
give inorganic sulphate and a fatty alcohol. The fatty
alcohol is oxidised to an aldehyde and subsequently
to a fatty acid with further oxidation via the L-oxidation pathway. AS and their degradation products
are ultimately biodegradable, Thomas and White [64]
observed that of 14 C SDS 70% was degraded to CO2
and all the remaining 30% was incorporated into the
microbial biomass, i.e. 100% of the SDS was utilised
for either energy or biomass production. Under anaerobic conditions digestion is rapid, as none of the
degradation pathways requires the presence of molecular oxygen. Swisher [10] reported s 90% removal
of AS in anaerobic test system. Steber et al. [63]
reported s 90% release of 14 C labelled AS as CO2
and CH4 . Birch et al. [48] reported values of 88% for
a similar experiment involving degradation of stearyl
sulphate. AS appear to be readily bioavailable under
both aerobic and anaerobic conditions and easily
degradable both primarily and ultimately. Therefore,
treatment in a WWTP is entirely su¤cient to eliminate AS and little possibility exists for these surfactants to reach the environment via sludge amendment.
3.6. Alcohol ether sulphates
Alcohol ether sulphates (AES) degrade well under
aerobic conditions with comparable primary and ultimate degradation rates to AS. Fischer [65] reported
primary degradation rates of 96% in 30 days in a
closed bottle test; Scho«berl et al. [55] reported ultimate degradation rates of 98^99% in the OECD
screening test. Steber and Berger [54] suggested three
possible degradation pathways for AES. (i) g/L-oxidation of the alkyl chain; (ii) cleavage of the sulphate
bond; and (iii) cleavage of an ether bond. Swisher
[57] stated that all three mechanisms are realised.
Hales et al. [66] suggested that ether cleavage is the
predominant mechanism but all three have been
shown to occur [66^68]. Very little data have been
published concerning the fate of AES in the anaerobic environment. Examination of the degradation
pathways suggests that cleavage of the sulphate
bond and ether bonds is possible without the pres-

245

ence of molecular oxygen. Oba et al. [69] reported
signi¢cant primary degradation of AES when anaerobically digested. Painter [43] reported high AES primary degradation in an anaerobic digester and a
septic tank over a period of 6^8 months. Itoh et al.
[70] revealed CO2 and CH4 production from AES
digestion, suggesting ultimate biodegradation if at a
somewhat reduced rate. Such data suggest that AES
are readily bioavailable in both aerobic and anaerobic environments.
3.7. Cationic surfactants
Cationic surfactants having a positive charge have
a strong a¤nity for the surface of particulates in
sewage sludge, which are predominantly negatively
charged. Kupfer [71] and Topping and Waters [72]
observed that in activated sludge 95% of the cationic
surfactants were adsorbed to the surface of particulate matter. Huber [73] observed that 20^40% cationic surfactants in primary settling tanks were associated with particulate matter. Cationic surfactants are
considered very biologically available [74]. Games et
al. [75] reported half-lives of 2.5 h for octadecyltrimethylammonium chloride in wastewater. Krzeminski et al. [76] reported alkylbenzyldimethylammonium chloride to be ultimately biodegradable with
s 80% of the 14 C labelled compound being released
as 14 CO2 . Sullivan [77] stated that ditallowdimethylammonium chloride (DTDMAC) in activated sludge
was predominantly associated with the particulate
matter and after digestion 40% was released as
14
CO2 . The adsorption of cationic surfactants to particulate matter increases the importance of understanding the anaerobic degradation processes. However, literature concerning cationic surfactant fate
under anaerobic conditions is scarce. Janicke and
Hilge [78] reported that quaternary ammonium salts
exhibited little or no degradation under anaerobic
conditions. Battersby and Wilson [79] observed that
concentrations of 200 mg l31 hexadecyltrimethylammonium bromide inhibited the production of methane, suggesting that such concentrations are inhibitory to the resident microbes. However, van Ginkel
[74] states that at the concentrations that cationic
surfactants are found in raw sewage there appears
to be no e¡ect upon wastewater treatment processes.
Aerobic degradation pathways for quaternary am-

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M.J. Scott, M.N. Jones / Biochimica et Biophysica Acta 1508 (2000) 235^251

monium salts are suggested in van Ginkel [74]. However, no pathways for anaerobic degradation exist
within the literature. The initial oxidation of the surfactant cannot take place without the presence of
molecular oxygen. Therefore, it can be assumed
that cationic surfactants are not anaerobically biodegradable either because of a lack of appropriate metabolic pathways and/or a possible toxic e¡ect of the
surfactant upon the relevant anaerobic microorganisms.

rivatives are present in relatively high concentrations
in sludge and can enter the environment via application to agricultural land. However, Marcomini et al.
[53] observed that sewage sludge amended soil exhibited a rapid drop in NP concentration post application with approximately 80% degradation within 3
weeks, suggesting that within the aerobic soil environment NP will not accumulate.

3.8. Alkyl phenol ethoxylates

Fatty alcohol ethoxylates (AE) were developed as
an eco-friendly alternative to APE. A great deal of
literature exists upon the biodegradability of these
compounds. Linear AE are considered readily biodegradable, Kravetz et al. [84] observed s 80% primary degradation in 28 days for linear AE and
40% for branched AE. Balson and Felix [80] suggest
that the breakdown mechanism of these compounds
involves the initial hydrophobe^hydrophile scission
of the AE yielding a hydrophobe and a polyalkoxylate, e¡ectively achieving primary degradation. The
hydrophobe then undergoes g/L-oxidation. Klotz
[85] reported AE concentrations in sludge of 6 700
mg kg31 . Such concentrations suggest that AE is not
entirely degradable under anaerobic conditions.
However, present understanding is somewhat limited
and too sparse to conclude such a fact. Knaebel et al.
[86] showed linear alcohol ethoxylates (LAE) to be
readily bioavailable in a variety of di¡erent soil
types, suggesting that aerobic soil amended with
sludge rich in AE will not accumulate the surfactant.

APE undergo almost complete primary degradation in the presence of oxygen [80]. Jones and Westmoorland [81] observed nonyl phenol ethoxylate
(NPE) degradation in composted sludge collected
from wool scouring. They observed a 98% net primary degradation of NPE in 100 days. g/L-oxidation
was attributed to the degradation of the alkyl chain,
but little evidence was observed of any degradation
of the aromatic ether bond. This observation was
supported by the build up of NP towards the end
of the monitoring period. Kravetz et al. [82] observed
a similar di¡erence in the degradation rates of the
breakdown products of APE. Radiolabelled APE
(3 H labelled aromatic constituent; 14 C labelled alkyl
chain) was placed in a bioreactor and terminal degradation products monitored. Twenty nine percent of
the 3 H label was converted to water but 58% of the
14
C was converted to CO2 . This suggests that though
rapid primary degradation takes place, degradation
products are not as available to microorganisms. The
polyoxyethylene chain appears to be readily biodegradable but the NP derivative appears more resilient. Due to their amphiphilic nature APE and their
breakdown products show an a¤nity for particulate
surfaces, and a signi¢cant proportion is observed
within the sludge fraction. Concentrations of APE
reported in the literature appear to be much higher
in anaerobically digested sludge (900^1100 mg
kg31 [53]) than in aerobically digested sludge (0.3
mg kg31 [83]). The degradation of APE and their
breakdown products appears restricted in the anaerobic environment. Without the presence of molecular
oxygen the initial g-oxidation of the alkyl chain cannot take place, restricting breakdown. Therefore, elevated concentrations of APE and its breakdown de-

3.9. Fatty alcohol ethoxylates

4. Surfactants in wastewater
Surfactants can reach the environment as a result
of discharge from WWTP into rivers and estuaries or
by direct discharge of raw sewage. Raw sewage discharge is increasingly rare in most industrial nations
although small amounts are still disposed of in this
way. The fate of all organic pollutants in WWTP is
determined by several processes including gas exchange with the atmosphere, sorption to suspended
solids and aerobic and anaerobic biodegradation
[87]. E¤cient treatment in WWPT will result in discharge of very low levels of surfactants into the environment. For LAS, SAS and the cationic surfac-

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Fig. 5. The decrease in water concentration of LAS in the river
Llobregat as a result of biodegradation/bioadsorption after the
discharge of raw sewage from Caserres (Barcelona, Spain) as a
function of distance from the discharge point.

tant DTDMAC sorption to sewage sludge accounts
for 26, 16 and 23 to 53% respectively of the in£uent
materials, while biodegradation removes 73, 43 and
36 to 43% respectively. The e¥uent from the WWTP
leaves only 1% LAS, 6 1% SAS and 6^41%
DTDMAC. The surfactants will undergo further biodegradation in the environment that together with
dilution will reduce their toxicological e¡ects further.
Studies of particular WWTP in Germany have
been reported by Schro«der et al. [88,89]. These cover
discharge from WWTP on the rivers Rur [88] and
Angerbach, a small tributary in the vicinity of Dusseldorf [89]. In the latter study the WWTP served a
population of approximately 59 000 plus an industrial equivalent of 10 000 and the total plant capacity
was 25 000 m3 of water. The LAS concentrations in
the in£uent peaked during daytime at around 19.00 h
at approximately 3500 Wg l31 , for AES the peak
was 4500 Wg l31 and for alcohol sulphates (AS) 600
Wg l31 . Monitoring of the surfactant in£uent and
e¥uent concentrations enabled elimination rates to
be calculated for anionic surfactants. These were
found to be 99.7% (LAS), 99.9% (AS) and 99.99%
(AES).

247

It is clear from these studies [88,89] and the data
reported by Alder et al. [87] for various WWTP that
anionic surfactant elimination is highly e¤cient in
modern WWTP, but somewhat less so for cationic
surfactants. This should not however, be cause for
complacency, the removal of constituents in detergent formulations such as £uorescence whitening
agents, naphthalene sulphates used in chemical,
pharmaceutical and textile industries and also the
organotin compounds in marine antifouling paints
such as tributyltin, is much less e¤cient. In the
case of naphthalene sulphates 95% of these pollutants is still present in WWTP e¥uents [87].
Direct discharge of raw sewage is becoming increasingly rare, but a study of the Llobregat river
near Barcelona (Spain) where raw sewage from Caserres is discharged, showed that the rate of biodegradation of LAS was rapid provided the £ow conditions in the river were adequate [90]. Fig. 5 shows the
decrease in concentration of LAS (on a log scale) as
a function of the distance from the discharge point of
the sewage. There is a rapid decrease in concentrations of LAS in the river water by 1.5 km from the
discharge point particularly in winter months when
the water £ow in the river is high (75 m3 min31 ). In
summer months (£ow rate 4.5 m3 min31 ) the decrease is less marked. At 4.8 km from the discharge
point the concentrations are similar in both seasons
and correspond to approximately 0.06% of the discharge concentration. It should be noted that the
decreases in concentration relate to both biodegradation and the loss due to adsorption on river sediments and suspended solids in the raw sewage.
The concentrations of LAS in surface water of the
North Sea, between 1 and 70 km from the coast,
were measured by Proctor and Gamble in February
1989. The results together with those of further measurements, starting from the mouth of the river
Scheldt up to 15 km o¡shore, made in October
1989 were reported in 1991 [91]. The LAS concentrations in these studies ranged from 6 0.05 Wg l31 to
9.4 Wg l31 . The salinity of the water is a major factor
controlling the LAS concentration in the marine environment. LAS adsorbs on river sediments in estuaries and the settling rates of sediments increase with
salinity when the river water mixes with the seawater.
The decreases in surfactant levels due to salinity are
much greater than predicted due to dilution. Stal-

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M.J. Scott, M.N. Jones / Biochimica et Biophysica Acta 1508 (2000) 235^251

mans et al. [91] conclude that typical LAS concentrations in the North Sea will be lower than 1 Wg l31 .
Taken in the context of margins of safety to marine
organisms, the LC50 values for various species are as
follows: cod (Gadus morrhua) 1 mg l31 , £ounder
(Platichthys £esus) 1.5 mg l31 , plaice (Pleuronectes
platessa) 5 mg l31 , oyster 0.025 mg l31 and pink
shrimp 19^154 mg l31 , even for the most sensitive
organisms (e.g. oyster) the safety margin is 25.
5. Conclusions
Examination of the data available suggests that
raw sewage passing through a modern WWTP has
a signi¢cant proportion of its surfactant load removed. Aerobic treatment processes appear to provide ideal conditions for rapid primary and ultimate
biodegradation via a consortium of bacteria. Wastewater e¥uent released into the environment appears
to have had its surfactant load reduced to an extent
that lethal e¡ects on aquatic organisms are negligible. Safety margins exist in excess of 25 for a variety of organisms. Problems associated with degradation products of APE, particularly NP, have led in
recent years to a reduction in APE usage with an
agenda towards eventual replacement by more ultimately degradable substitutes. Meanwhile it is acknowledged that release of the more recalcitrant
NP may be implicated in the problems experienced
from environmental oestrogen mimics.
From the data available it is evident that LAS,
SAS, FES, cationic surfactants, APE and AE are
all relatively resistant to degradation in anaerobic
environments. As anaerobic digestion is the predominant treatment of sludge from primary settling
tanks, and the amphiphilic nature of surfactants promotes their adsorption to particle surfaces in sewage,
there appears an opportunity for surfactants to pass
through a WWTP relatively untreated. Application
of sludge to agricultural land may provide a large
source of surfactants to the soil environment. However, it appears that once re-introduced into an aerobic environment, such as soil, the surfactants are
once again readily bioavailable. From studies on
LAS, APE (and its derivatives) and FES, it can be
concluded that some surfactants are not anaerobically degradable during sludge treatment, but are rap-

idly degraded when applied to aerobic soils. In the
literature it is suggested that SAS and AE will be
readily available in the aerobic soil environment,
however there is a severe lack of relevant ¢eld studies
available to support such a statement. Of greater
worry is the lack of data concerning the degradation
of cationic surfactants and the oestrogen mimicking
properties of APE. Cationic surfactants, though
readily biodegradable in aerobic environments, are
toxic even at low concentrations. Therefore application to agricultural soil may have detrimental e¡ect
to the soil biota. APE is readily primarily degradable
aerobically, however NP, one of the primary degradation products has been implicated as an environmental oestrogenic compound. NP has a strong af¢nity for soil particles and is less biodegradable than
APE. At present the authors could ¢nd little published literature concerning the oestrogen mimicking
properties of NP in sludge amended soils, this is a
¢eld that needs further research. Also failure of
WWTP to remove more e¡ectively other constituents
present in wastewater arising from detergents (builders, whiteners, blueing and bleaching agents, etc.) is a
cause of some concern.
Three major routes of sewage sludge disposal exist
in the UK, application to soil, land¢ll and incineration. Pressure from environmentalists and rising
costs are gradually making land¢ll and incineration
less attractive. This combined with increasing
amounts of sludge produced, is increasing the pressure on agricultural application. It appears that surfactant application to aerobic soils is quite safe due
to rapid biodegradation rates. However, the temptation to dispose of sludge on non-agriculture soils
should be carefully investigated. Soils that are anaerobic may not be appropriate sites for amendment.
Such soils may exhibit accumulation of surfactants
as biodegradation is retarded and may result in surfactant contamination of the environment.
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