Anaerobic Biodegradation of Surfactants – Scientific Review5.4 Wastewater treatment 5.5 Landfill 5.6 Animals 6 Existing Methods 319 6.1 Screening tests 6.1.1 Anaerobic screening

y J. L. Berna, G. Cassani, C.-D. Hager, N. Rehman, I. López, D. Schowanek, J. Steber, K. Taeger and T. Wind

Anaerobic Biodegradation of Surfactants –Scientific Review

The anaerobic biodegradation of surfactants is used as an ac-ceptability criterion in some environmental pieces of legislation(eco-label, risk assessment, etc.), without a proper evaluation ofthe relevance of such a characteristic. Available screening testmethods to assess the anaerobic biodegradation do not simu-late the real conditions prevailing in these anaerobic compart-ments but rather reflect more stringent conditions, due to thehigh test substance/biomass ratio, possibility of inhibitory ef-fects and limited possibility for adaptation. Therefore positiveresults are indicative of a similar behaviour under environmen-tal conditions, while a negative result cannot be necessarily in-terpreted as inherent anaerobic recalcitrance. The majority ofsurfactants entering the environment will be exposed to anddegraded under aerobic conditions, and only less than 20%will potentially reach temporarily anaerobic environmentalcompartments. In contrast to the well documented adverse ef-fects observed in the absence of aerobic biodegradation, thelack of anaerobic biodegradation does not seem to be corre-lated with any apparent environmental problem for most com-partments after more than 40 years of widely use of suchproducts. The scientific review concluded that anaerobic biode-gradability does not have the same environmental relevance asthe aerobic one. Anaerobic biodegradability should not, there-fore, be used as a pass/fail property for the environmental ac-ceptability of surfactants which are readily biodegradable underaerobic conditions.

Key words: Surfactant, anaerobic biodegradation, anaerobicbiodegradability, wastewater treatment, screening tests, simula-tion tests, sediment, monitoring data

Anaerober Abbau von Tensiden – Wissenschaftliche Über-sicht. Der anaerobe Abbau von Tensiden wird häufig – ohnewissenschaftliche Bewertung der Relevanz einer solchenKenngröße – als ein Akzeptanzkriterium in einigen umweltrele-vanten Teilen von Verordnungen (Ecolabel, Risikoabschätzun-gen, etc.) genutzt. Bekannte Screening-Testmethoden zur Ab-schätzung des anaeroben Abbaus simulieren nicht realeVerhältnisse, die in anaeroben Kompartimenten vorherrschen,sondern spiegeln aufgrund des hohen Testsubstanz/Biomasse-Verhältnisses sowie möglicher Inhibitionseffekte und begrenztmöglicher Adaption deutlich strengere Bedingungen wider.Folglich weisen in Screening-Tests positive Abbauergebnisse aufein ähnliches Verhalten unter Umweltbedingungen hin, wäh-rend ein negatives Ergebnis nicht unbedingt als inhärent anae-robe „Hartnäckigkeit“ zu interpretieren ist. Die Mehrzahl der indie Umwelt eingetragenen Tenside wird aeroben Bedingungenexponiert und unter diesen auch abgebaut. Nur weniger als20% der Tenside erreichen eventuell temporär anaerobe Um-weltkompartimente. Im Gegensatz zu gut dokumentierten nega-tiven Effekten, die in Ermangelung von aerobem Abbau beo-bachtet werden, scheint für die meisten Kompartimente einmangelnder anaerober Abbau – auch nach mehr als 40 Jahrenweitem Gebrauch entsprechender Tenside – nicht mit irgendei-nem ersichtlichen Umweltproblem assoziiert werden zu kön-

nen. Dieser wissenschaftliche Übersichtsartikel schlussfolgert,dass die anaerobe Abbaubarkeit nicht von gleicher Relevanz ein-zustufen ist wie die aerobe. Daher sollte die anaerobe Abbau-barkeit nicht als ein Entscheidungskriterium zur Umweltakzep-tanz von Tensiden genutzt werden, die unter aerobenBedingungen leicht biologisch abbaubar sind.

Stichwörter: Tenside, anaerober Abbau, anaerobe Abbaubar-keit, Abwasserbehandlung, Screening-Tests, Simulationstests,Sediment, Daten-Monitoring

Table of Contents

Foreword and Position Paper 313

1 Introduction 314

2 Executive Summary 314Environmental relevance of anaerobic biodegrad-ability of surfactants

3 Surfactants in Europe Production/Consumption 317

4 Definitions 317Anaerobic biodegradation

5 Anaerobic Compartments 3185.1 Overview5.2 Terrestrial5.3 Aquatic

5.3.1 Water5.3.2 Sediment5.3.3 ‘Extreme’ habitats5.3.4 Groundwater

5.4 Wastewater treatment5.5 Landfill5.6 Animals

6 Existing Methods 3196.1 Screening tests

6.1.1 Anaerobic screening tests based on gasvolume measurement only

6.1.2 Anaerobic screening tests based on gasproduction measurement in the gas andthe liquid phase

6.1.3 Predictive value of anaerobic biodegrada-tion screening test data

6.2 Simulation tests6.2.1 Introduction to anaerobic simulation

tests6.2.2 Test systems

6.2.2.1 14C-Anaerobic digester simulationtest

6.2.2.2 Anaerobic transformation test inaquatic sediment systems(OECD guideline 308)

6.2.2.3 Continuously operated anaerobicreactors

REVIEW ARTICLE

312 ª Carl Hanser Publisher, Munich Tenside Surf. Det . 44 (2007) 6

6.2.2.4 Biological nutrient removal STPsimulation tests: Behr and CAS-UCT units

6.2.2.5 Others systems

7 Interpretation of Available Data on AnaerobicBiodegradation of Surfactants 3247.1 Anionics

7.1.1 Sulfonates7.1.2 Sulfates7.1.3 Fatty acids and soaps

7.2 Nonionics7.3 Cationics – Amphoterics

8 Evaluation of the Relevance of Anaerobic Biodegra-dation 3348.1 Understanding mass fluxes of surfactants in

environmental compartments8.1.1 Mass fluxes obtained via modelling for

hypothetical surfactants8.1.2 Mass fluxes based on monitoring data

8.2 Impact of surfactants on structure and func-tion of anaerobic environmental compart-ments8.2.1 Speciation and bioavailability of surfac-

tants in anaerobic compartments8.2.2 Impact on wastewater treatment and

aquatic environments8.2.2.1 LAS and other anionic surfactants

in anaerobic digesters8.2.2.2 Other surfactants in anaerobic di-

gesters8.2.2.3 Septic tanks and decentralized

treatment systems8.2.2.4 Anaerobic river and lake sedi-

ments8.2.2.5 Agricultural soil (pasture and

cropland)8.2.2.6 Landfills for sludge8.2.2.7 Marine sediments

References 340

Glossary Of Abbreviations 343

Appendix 344Overview of surfactant monitoring and mass balan-cing in anaerobic environmental compartments

Foreword to Revision 0 (1999)

This document was commissioned by ERASM (Environ-mental Risk Assessment Steering Committee – a detergentindustry group).

ERASM was created in 1991 as a forum for the co-ordi-nation of views and actions in the field of risk assessmentand risk assessment issues between the chemical industry(CESIO – Comit� Europ�en des Agents de Surface et leursInterm�diaires Organiques, a CEFIC Sector Group) and thedetergent industry (AISE, Association Internationale de laSavonnerie, de la D�tergence et des Produits d’Entretien).

ERASM’s mission is to define the industry position inestablishing and managing the risk of surfactants to the en-vironment and to propose reasonable and technically soundguidelines for assessing the environmental risk of surfac-tants. ERASM also ensures that resources are available tocarry out necessary research programmes.

The structure of ERASM consists of ad hoc task forcesestablished to work on specific topics related to risk assess-

ment methods and data. A task force was set up to addressanaerobic biodegradation and the environmental relevanceof anaerobic biodegradability of surfactants.

The present dossier gives a compilation and interpreta-tion of the available literature on the fate and biodegradabil-ity of commercial surfactants under anaerobic conditions.

The members of this specific task force are:

José Luis Berna (Chairman)Nigel BattersbyLuciano CavalliRichard FletcherAndreas Guldner

Diederik SchowanekJosef Steber

Petresa (Es)Shell Chemicals (UK)Condea Augusta (It)

Unilever (UK)BASF (De)

Procter & Gamble (Be)Henkel (De)

Foreword to Revision 1 (2007)

The document published in 1999 has been updated withnew data published in the scientific literature in the period1999–2005. The new dossier represents therefore the mostrecent compilation of data, interpretation and assessmentof the relevance of anaerobic biodegradation of surfactants.

The report also contains data referring to surfactantswhich are not considered as relevant in the detergent indus-try today. They have been included however with the pur-pose of having a comprehensive data base.

The members of the ERASM task force responsible forthis update are:

y Jos� Luis Berna (Chairman) Petresa (Spain)y Giorgio Cassani Sasol (Italy)y Claus-Dierk Hager Sasol (Germany)y Ignacio L�pez Serrano Petresa (Spain)y Diederik Schowanek Procter and Gamble

(Belgium)y Josef Steber Henkel (Germany)y Thorsten Wind Henkel (Germany)y Klaus Taeger BASF (Germany)y Naheed Rehman Unilever (UK)

Position Paper

The anaerobic biodegradation of surfactants is used as an ac-ceptability criterion in some environmental pieces of legisla-tion (eco-label, risk assessment, etc.), without a proper eval-uation of the relevance of such a characteristic.

Surfactants form a group of chemicals with considerableenvironmental importance due to their high volume con-sumption and widespread use as they are essential ingredi-ents in most laundry and cleaning products. Since the majorpart of the biosphere is aerobic, priority has been given tothe study and assessment of biodegradability under theseconditions. Nevertheless there are environmental compart-ments which can be permanently (e.g. anaerobic digesters)or temporarily anaerobic (e.g. river sediments and soils) andsurfactants do reach these.

Available screening test methods to assess anaerobic bio-degradation do not simulate the real conditions prevailing inthese anaerobic compartments but rather reflect more strin-gent conditions, due to the high test substance/biomas ratio,possibility of inhibitory effects and limited possibility foradaptation. Therefore positive results are indicative of a sim-ilar behaviour under environmental conditions, while a ne-gative result cannot be necessarily interpreted as inherentanaerobic recalcitrance. In addition, low biodegradation re-sults in these tests may be influenced by a limited bioavail-ability due to the formation of insoluble chemical species.

J. L. Berna et al.: Anaerobic biodegradation of surfactants – scientific review

Tenside Surf. Det . 44 (2007) 6 313

The relevance of anaerobic biodegradability cannot beseparated from other important properties of surfactantssuch as sorptive behaviour, ecotoxicity profile and above all,aerobic biodegradation rate.

The majority of surfactants entering the environmentwill be exposed to and degraded under aerobic conditions,and only less than 20% will potentially reach anaerobic, en-vironmental compartments. In all but a few cases their pre-sence in these will not be permanent.

A systematic evaluation of the risk to the structure andfunction of these compartments due to the presence ofnon-degraded surfactants led to the conclusion that, in con-trast to the adverse effects observed in the absence of aerobicbiodegradation, the lack of anaerobic biodegradation doesnot seem to be correlated with any apparent environmentalproblem for most compartments. Particularly for the sedi-ment compartment, data is lacking and it is recommendedto fill the missing data gaps to assess structure and function.

In criteria for eco-labelling a conservative set of “scoring”or “weighting” factors, if any, for anaerobic biodegradability,should follow from a combination of the above characteris-tics, and it is suggested that these should be of the order ofone tenth of the aerobic biodegradability value for readilybiodegradable surfactant.

Consequently it is concluded that anaerobic biodegrad-ability does not have the same environmental relevance asthe aerobic one. Anaerobic biodegradability should not,therefore, be used as a pass/fail property for the environ-mental acceptability of surfactants which are readily biode-gradable under aerobic conditions.

1 Introduction

Biodegradation is the most important mechanism for theirreversible removal of chemicals from the aquatic and ter-restrial environments. Therefore, the evaluation of biode-gradability is an indispensable element of the exposureassessment within the environmental risk assessment pro-cess of a chemical substance.

The biosphere is predominantly aerobic and it is under-standable that the biodegradation behaviour of chemicalsunder aerobic conditions has been the focus of attentionfor a long time. This has led to the development of a num-ber of internationally used and recognised laboratory meth-ods for assessing aerobic biodegradability and to a hugeamount of test data. The biodegradability of chemicals inthe absence of free oxygen, i. e., under anaerobic conditionshas been considered to a far lesser extent. Nevertheless,there are environmental areas which are permanently ortemporarily anaerobic (e.g. sludge digesters of sewage treat-ment plants, sediments or sub-surface soil layers). It can beargued therefore that a comprehensive evaluation of a chem-ical’s environmental fate should also address its anaerobicbiodegradability, provided that there is a possibility for enter-ing anaerobic environments to a significant extent.

Typically, such substances are characterised by a poorwater solubility or strong adsorptivity onto solids, resultingin an environmental distribution with a pronounced trans-portation into anaerobic compartments. Surfactants form agroup of chemicals with a high environmental relevanceand physical-chemical properties which may result in a sig-nificant partitioning between the aqueous and the solidphase in the aquatic environment. Environmental monitor-ing data for some surfactants support this partitioning as-sumption, showing their presence in secondary and digestersludges, and sediments. The same substance properties arealso relevant in the context of the environmental risk assess-

ment of chemicals in the soil compartment as application ofdigested sludges for agricultural purposes will influence theinitial environmental concentration in soils.

In order to establish an overview of the anaerobic biode-gradation aspects in general (i. e. environmental relevance,test methodology) and of their relevance for surfactants inparticular, an ERASM Task Force was created. This positionpaper was prepared by experts of the involved industries andaims to form a common basis of knowledge on this topic.

2 Executive Summary

Environmental relevance of anaerobic biodegradabilityof surfactants

Aim of the review

The objective of this review, prepared by the ERASM TaskForce ‘Anaerobic Biodegradability’ was to provide a thoroughcompilation and interpretation of the available literature onthe fate and biodegradability of commercial surfactants un-der anaerobic conditions. This analysis should form the ba-sis for a data-based discussion on the environmental rele-vance of the property of anaerobic biodegradability (or lackthereof) in a broad range of contexts, such as risk assess-ment of chemicals, the upcoming review of the DetergentsRegulation (EC) 907/2004 (addressing specifically the anae-robic biodegradability of detergent surfactants, the biode-gradability of non surfactant organic ingredients and theuse of phosphates) and the eco-labelling of detergent pro-ducts.

The earth’s atmosphere contains just over 20% of oxygenby volume making aerobic compartments the predominantones in the biosphere. It is therefore not surprising that theprevalent route for biodegradation of chemicals is aerobicand, hence, aerobic biodegradation has been extensivelystudied. In addition, it is a legal requirement in the Eu-ropean Union (Detergent Regulation) for surfactants usedin detergents to be readily (and ultimately) biodegradableunder aerobic conditions.

Surface active agents are essential ingredients in mosthousehold laundry products, domestic and industrial clean-ers, as well as in personal care, cosmetic products and a vari-ety of industrial applications. Surfactants form a group ofchemicals with a high overall environmental relevance, dueto a combination of their inherent environmental propertiesand their very large production volume. They are typicallydischarged into the environment through the sewage treat-ment infrastructure (i. e. sewers, sewage treatment plants),or directly in situations where no treatment systems areavailable.

Anaerobic test methods and their predictive value

Test methods to determine the ultimate anaerobic biode-gradability of organic compounds at screening and simula-tion level are available. First, there are screening tests suchas the ECETOC test/ISO 11734/OECD-311 which deter-mine anaerobic mineralization by measurement of themethane and carbon dioxide gas production via a pressureor volumetric reading and DIC measurement. Due to thestringent screening-type test conditions (i. e. high test sub-stance/biomass ratio), inhibitory effects of surfactants can-not be excluded and the possibilities for acclimation are lim-ited. These tests are reliable at avoiding false-positiveconclusions, but a poor result is not necessarily proof ofanaerobic recalcitrance. In simulation tests the substance/biomass ratio is much lower and can reach real conditions


314 Tenside Surf. Det . 44 (2007) 6



(e.g. 8,000 mg C/kg dry solids, if gas production is evalu-ated, or even lower, if specific analytics is applied).

Anaerobic biodegradation test systems under more rea-listic conditions (in terms of sludge concentration, freshsludge feed, residence times, etc.) include digester simula-tion tests employing radio labelled test substances or discon-tinuous systems. Limitations, however, remain for situationssuch as low substrate concentrations and toxic surfactants.While, in principle, test methods for simulation of aerobic/anaerobic river sediments exist, experimental data based onthese improved testing methodologies are scarce. Also, theapplication of specific analytical methods to follow the fateof surfactants and their breakdown products in anaerobicmatrices is not widespread as yet.

Anaerobic biodegradability of surfactant groups

Based on the aforementioned laboratory test methods, theavailable data from published biodegradation studies andavailable monitoring studies allow the evaluation of theanaerobic ultimate biodegradability of the important surfac-tant groups used in detergents:

y Sulfonated anionic surfactants (LAS, SAS, MES, AOS):poorly biodegradable

y Sulfated anionic surfactants (AS, AES): well biodegrad-able

y Fatty acids and soaps: well biodegradabley Alcohol ethoxylates (AE): well biodegradabley Sugar-based non-ionic surfactants (APG, glucamides):

well biodegradabley Alkylethanolamides: well biodegradable (based on lim-

ited data)y Alkyldimethyl amine oxides: well biodegradable (based

on limited data).y APEOs*: partially degradable leaving alkylphenol resi-

dues.y Mono- or di-alkyl quaternary compounds (TMAC,

DTDMAC*): poorly biodegradable.y Esterified mono- or di-alkyl quaternary surfactants (ester-

quats): well biodegradable.y Amphoteric surfactants (alkylamido betaines, alkylbe-

taines, alkylamphoacetates): well biodegradable (basedon limited data)

* Not relevant anymore in detergent applications

Factors for evaluating the importance of anaerobic biodegrada-tion

Mass flux to anaerobic environments: Strictly aerobic andanaerobic environments represent the two extremes of acontinuous spectrum of environmental habitats which arepopulated by a broad variety of micro-organisms with specif-ic biodegradative capabilities. Due to the specific usage pat-terns, the bulk of the surfactant mass entering the environ-ment will predominantly be exposed to and degraded underaerobic conditions. For readily biodegradable surfactants,less than 1% of the flux will reside in the permanently anae-robic compartments of a landfill and the occasionally anae-robic compartment of agricultural land.

The relative importance of sewage treatment has in-creased considerably during the last decade with a highertreatment incidence, halting of sludge disposal into the seaand improved treatment systems with biological nutrient re-moval. Incineration of waste water sludge has become amore important disposal route. Consequently, the amountof surfactants which can reach anaerobic compartments viasludge disposal has decreased and is likely to decrease

further over time (Urban Wastewater Treatment Directive91/271). In spite of the EU Directive 91/271 generally stillthe quality of wastewater treatment is very different in theEU.

The differences in physico-chemical properties and bio-degradability explain the observed differences between themass fluxes of different surfactants. As such, the relevanceof anaerobic biodegradability cannot be separated fromother important properties of the surfactants, such as sorp-tive behaviour (log Kow), solubility, and most importantly,aerobic biodegradation kinetics.

Accordingly, a German Expert Group for Detergents(HAD)1 had previously proposed ranking the relevance ofanaerobic biodegradability as a function of the above proper-ties.

The recently conducted environmental risk assessmentsof all important surfactants used in detergents (HERA) un-derline the considerably higher environmental relevance ofthe aerobic biodegradability. Recently (Nov. 2005), SCHERopinion on the “Environmental Risk Assessment of nonBiodegradable Detergent Surfactants under Anaerobic Con-ditions” has concluded that the requirement for ready and ulti-mate anaerobic biodegradability under anaerobic conditions isnot by itself regarded as an effective measure for environmentalprotection. These (aerobically) readily biodegradable sub-stances proved to be of no concern for the aquatic and ter-restrial environment regardless of their anaerobic biodegrad-ability.

Bioavailability:Changes of bioavailability of surfactants under anaerobicconditions can affect the outcome of toxicity/inhibition stud-ies, and to some extent biodegradation or removal rates. Thespecific chemical structure of some surfactants contributesto a rapid precipitation with water hardness ions (Ca++,Mg++) into insoluble forms, as well as adsorption to the sur-rounding solid matrix. This highlights the need to use thereal environmental form of a surfactant in inhibition andbiodegradation tests, in order to obtain a realistic testresult.

Impact of surfactants on the structure and function of anaerobiccompartments:Based on available data, an evaluation was made of the riskto the structure and function of anaerobic compartmentsbeing affected by fluxes of non-degraded surfactants enter-ing the system. The evaluation led to the conclusion that incontrast to the adverse effects and high risk in the absenceof aerobic biodegradation, the lack of anaerobic biodegrada-tion does not seem to be correlated with any apparent envir-onmental or technical problems in most compartments.Nevertheless, it is fair to say that for natural anaerobic envir-onments, not all aspects of structure and function can beadequately assessed today since comprehensive data arelacking. Because of the precautionary principle knowledgeof the anaerobic degradability of surfactants would be pre-cious. It is acknowledged therefore that surfactants biode-gradable under both aerobic and anaerobic conditions leaveless room open to environmental issues.

Conclusions

Anaerobic biodegradability as a strict pass/fail criterion isnot in line with the environmental interpretation and signif-

1 Schöberl P. (1994). Die Bedeutung fehlender anaerober biologischer Abbau-barkeit. Tenside Surf. Det. 31,157–162.

icance that should be given to the lack of this property forsurfactants. For surfactants in detergents, rapid aerobic bio-degradation as well as their sorptive and ecotoxicologicalproperties are keys for a realistic environmental compatibil-ity assessment. If a surfactant is rapidly degradable underaerobic conditions, and its (transitory) presence in anaerobicenvironments does not affect the function and structure ofthat compartment (e. g. it is not inhibitory), then its anaero-bic degradability is of minor importance. Nevertheless, it isfair to say that for natural anaerobic environments, not allaspects of structure and function can be adequately assessedtoday since comprehensive data is lacking.

3 Surfactants in Europe Production/Consumption

Surfactants (anionics, nonionics, cationics and amphoterics)are used in different fields such as cosmetics, metal work-ing, mining, agriculture, paper and leather industries andobviously the detergent industries.

CESIO compiles annual statistics on the production andconsumption of surfactants in Europe.

The total surfactant market in Western Europe is matureand annual growth rates are expected to be 1–1.5%.

The largest market for surfactants is in household appli-cations, which accounts already for half of the total surfac-tant consumption. In this market a limited number of sur-factant types remain dominant: alkylbenzene sulfonates,alcohol-sulfates and alcohol ether sulfates, alkane sulfonatesas well as fatty alcohol ethoxylates. Other outlets for surfac-tants are I&I (Industrial and Institutional cleaning), perso-nal care, textiles, emulsion polymerization, paint additives,agrochemicals, etc. where besides the a.m. commodity typesurfactants, mainly “specialty” surfactants are used. Figure 1illustrates the wide field of applications of surfactants.

4 Definitions

Anaerobic biodegradation

Anaerobic biodegradation means the microbial degradationof organic substances in the absence of free oxygen (O2).While oxygen serves as the electron acceptor in aerobic biode-gradation processes forming water (H2O) as the final product,degradation processes in anaerobic systems depend on alter-native acceptors such as sulfate, nitrate or carbonate yielding,in the end, hydrogen sulphide (H2S), molecular nitrogen (N2)and/or ammonia (NH3) and methane (CH4), respectively.

Anaerobic biodegradation is a multistep process per-formed by different bacterial groups. It involves hydrolysisof polymeric substances like proteins or carbohydrates tomonomers and the subsequent decomposition to solubleacids, alcohols, molecular hydrogen (H2) and carbon dioxide(CO2). Depending on the prevailing environmental condi-tions, the final steps of ultimate anaerobic biodegradationare performed by denitrifying, sulfate-reducing or methano-genic bacteria (Figure 2).

In contrast to the strictly anaerobic sulfate-reducing andmethanogenic bacteria, the nitrate-reducing micro-organ-isms as well as many other decomposing bacteria are mostlyfacultative anaerobic (i. e. they are able to grow and to de-grade organic substances under aerobic as well as anaerobicconditions). Thus, strictly aerobic and anaerobic environ-ments represent the two extremes of a continuous spectrumof environmental habitats which are populated by a broad



Surfactants consumptionin Western Europe

(incl. captive use consumption)

in 1000 tonnes – as 100% active 1995 2000 2005

A. ANIONICS

Alkylbenzene sulfonates 410 450 490

Alkane sulfonates 74 72 64

Alcohol sulfates 94 86 60

Alcohol ether sulfates 192 261 400

Other anionics 80 75 90

Total Anionics (excl. soaps) 850 944 1104

Soaps 550 650 700

B. NONIONICS

Alkylphenol ethoxylates 82 86 30

Alcohol ethoxylates 440 470 530

Other ethoxylates 100 124 200

Other nonionics 170 163 120

Total Nonionics 792 843 880

C. CATIONICS 181 166 200

D. AMPHOTERICS 38 45 76

GRAND TOTAL 1861 1998 2260

Table 1 CESIO Statistics

Figure 1 Surfactant end uses in W. Europe

Figure 2 Anaerobic degradation processes by bacterial consortia

variety of micro-organisms with specific biodegradation abil-ities. Anaerobic conditions occur where vigorous decomposi-tion of organic matter and restricted aeration result in thedepletion of oxygen. Anoxic conditions may represent an in-termediate stage where oxygen supply is limited, still allow-ing a slow (aerobic) degradation of organic compounds. In adigester the various bacteria also have different require-ments to the surrounding environment. Acidogenic bacterianeed pH values from 4 to 6, whilst methanogenic bacteriafrom 7 to 7.5 (Temper and Pfeiffer, 1986). In batch tests thedynamic equilibrium is often interrupted because of an en-richment of acidogenic bacteria as a consequence of lackingsubstrate in- and outflow.

5 Anaerobic Compartments

5.1 Overview

The biosphere is predominantly aerobic with an estimated1015 tons O2 being present in the atmosphere and oceans(Press and Siever, 1986). However, anaerobic environmentscan develop where the consumption of O2 by the biologicaloxidation of organic matter exceeds supply. These may besmall anaerobic zones in an otherwise oxic system (e.g. Jør-gensen, 1977a), or much larger and stable environmentssuch as those found in marine and freshwater sediments.As was mentioned in Chapter 4, in the absence of free O2

the oxidation of organic matter by micro-organisms con-tinues through the use of a sequence of alternative electronacceptors (e.g. nitrate sulfate carbon dioxide). These reac-tions are dependent on the availability of organic and inor-ganic substrates, the redox potential of the environmentand the types of bacteria present (Zehnder and Stumm,1988). This is illustrated in Figure 3 and shows a progres-sion down idealised coastal sediment. It can be seen thatthe higher the redox potential of the environment, the moreenergetically favourable is the reaction. Hence, a facultativeanaerobic bacterium existing in an anaerobic zone with alow redox potential, will quickly become more energy effi-cient once there is an ingress of O2. This sequence of bacter-ia preferring reactions which provide the most energy yieldcan be found in a wide range of aquatic and terrestrial envir-onments.

Anaerobic processes play an important role in nutrientcycling and animal nutrition, and in man-made activitiessuch as wastewater treatment. This chapter summarisesthe main anaerobic environmental compartments.

5.2 Terrestrial

Soils are usually aerobic systems (Kaspar and Tiedje, 1982),although anaerobic micro sites can occur in poorly drainedsoils and a flooded soil such as a rice paddy may have anaerobic layer of only 1 cm depth (Richards, 1987). Anotherimportant terrestrial anaerobic environment is the peat inwater-logged moor land, fens and bogs.

5.3 Aquatic

5.3.1 Water

An anaerobic body of water can occur if there is high O2

consumption in the sediment or overlying water, and seaso-nal or permanent hydrographic conditions such as thermo-clines or haloclines prevent the exchange of water. Manyfreshwater lakes of >10 m depth can become stratified atthe end of the summer, leading to the formation of an anae-robic layer above the sediment (Jones, 1982).

The largest anaerobic ecosystems in the biosphere aremarine (Schlegel and Jannasch, 1981), with the Black Seabeing anaerobic from a depth of *200 m to the bottom at*2,000 m. Other permanently anaerobic basins are the Car-iaco Trench, off the coast of Venezuela (Richards, 1975) andthe Orca Basin in the northern Gulf of Mexico (Shokes et al.,1977). These reports and others have led Sieburth (1987) toconclude that anaerobic basins are fairly common. However,the decomposition of organic material in these regions ap-pears to occur predominantly in the oxygenated water col-umn, with reports of *75% of organic matter being aerobi-cally oxidised in stratified lakes (Rudd and Hamilton, 1979;Harrits and Hanson, 1980) and 85–95% in marine systemssuch as the Black Sea (Deuser, 1971).

5.3.2 Sediment

Depending on the level of eutrophication, water depth andseason, freshwater sediments are usually anaerobic belowthe surface few mm or cm. In general, aerobic degradationis the predominant decomposition process in these sedi-ments, although anaerobic processes such as dissimilatorynitrate reduction and methane evolution are also important(Jones, 1982).

The aerobic zone of marine sediments can vary from onlya fewmm in coastal areas to ‡1 m in deep sea sediments (Jør-gensen, 1982). The depth of this zone is dependent on factorssuch as the rate of O2 diffusion into the sediment porewater,bioturbation, microbial respiration rates (which in turn aredependent on organic carbon levels), sediment particle sizeand tidal flushing. Within the anaerobic layers, sulfate reduc-tion is the predominant terminal step in anaerobic biodegra-dation and this reflects the abundance of sulfate in sea water(29 mM at a salinity of 35%). In coastal sediments, themineralization of organic material by sulfate reduction is sig-nificant and has been reported to match or even exceed thatdue to aerobic respiration in intertidal salt marshes and mud-flats (Howarth and Teal, 1979; Mountfort et al., 1980) and fjor-dic sediments (Jørgensen, 1977b). However, these shelf (i. e.shoreline to 200 m depth) sediments cover only around 9%of the world’s oceans (Jørgensen, 1982) although they receivea disproportionately large amount of anthropogenic material.

5.3.3 ‘Extreme’ habitats

These are anaerobic environments which have extreme con-ditions of temperature (>100 �C), salinity (saturated NaCl) or



Figure 3 Simplified diagram of the different reactions performed by bacteriadepending on the redox potential of the environment. This example is forcoastal marine sediment and is not to scale

pH (<2, >10) and include geothermal vents, hot springs,salt lakes, peat bogs and alkaline hypersaline waters (seeLowe et al., 1993 for a detailed review). The release of surfac-tants into these habitats is virtually non-existent and as suchthey are ignored in this document.

5.3.4 Groundwater

Groundwater can become anaerobic, particularly when ashallow aquifer is contaminated with degradable organicmaterial from a landfill leachate (Smolenski and Suflita,1987). In deeper aquifers groundwater also can becomeanaerobic because of natural percolation of rain water andmicrobial activities. CO2 is absorbed during this passageand Mn2+ and Fe2+ are dissolved. Therefore not only con-tamination through organic material may lead to anaerobicconditions (Bavarian Environment Agency).

5.4 Wastewater treatment

Mesophilic (i. e. *35 �C) anaerobic digestion of sewagesludge is widely used in wastewater treatment to reduce vol-ume, stabilise the sludge and produce CH4-rich biogaswhich can be burnt to generate energy. Sludge digestion isa strictly anaerobic process, with a redox potential of the or-der of –250 mV (Mosey, 1985). Digester capacities usuallyrange from 1,000 to 10,000 m3 (Mosey, 1983) and the aver-age residence time is 15 to 20 days. Anaerobic digestionhas also been promoted as a means of supporting the fueldemands of developing countries, and large numbers of di-gesters have been installed in China and India (Compag-nion and Nyns, 1986).

Surfactants tend to adsorb onto sludge and as such arepassed into anaerobic digesters (Field et al., 1995). This ispossibly the anaerobic environment with the highest expo-sure to surfactants (see also the monitoring summary tablein the Appendix). In a number of European countries, thedigested sludge is eventually disposed to land as a soilamendment, whereby the plant and soil life may be exposedto surfactants or their metabolites.

Other anaerobic environments associated with waste-water treatment are: denitrifying filters and activated sludgesystems (Mosey, 1983), anaerobic ponds for high strengthindustrial wastewaters (Mara and Pearson, 1986), primitivetreatment plants based on septic tanks where wastewater ispercolated, and anaerobic/anoxic zones in wastewater treat-ment plants with biological removal (see below).

An increasing number of modern wastewater treatmentplants are fitted with anaerobic/low dissolved O2 zones inwhich enhanced nutrient (phosphorous, nitrogen) removaltakes place. Tertiary wastewater treatment partly includingsuch biological nutrient removal processes covers around75% of the population connected to wastewater treatmentplants in Northern Europe, ca. 64% in Central Europe and<10% in Southern Europe and the EU Accession Countries(EEA, 2002)2. The mentioned anaerobic/low dissolved O2

zones are usually situated prior to the aerobic zone and assuch surfactants may be subject to anaerobic/low dissolvedO2 attack prior to their aerobic biodegradation. The resi-dence time of the mixed liquor suspended solids in boththe anaerobic and the low dissolved O2 tank is limited toonly a few hours. Anaerobic degradation steps can bebrought about by the fraction of the biomass active in these

zones, which are mainly facultative anaerobic micro-organ-isms. Fermentation reactions will take place, although thefinal conversion to methane is not significant due to the lim-ited residence time in the anaerobic zones. Redox conditionsvary considerably over time and place in the system. Due tothe longer hydraulic and sludge residence times in such sys-tems and the diversity of degradation (redox) conditions, it isexpected that surfactant removal will generally be enhancedin comparison to classical activated sludge plants (Rottierset al., 1998).

5.5 Landfill

Decomposition of organic material in landfills, particularlythose with compacted deposits, leads to the consumption ofO2 and the formation of anaerobic conditions (K�ster andNiese, 1986).

5.6 Animals

Anaerobic environments exist in the alimentary tracts ofmany animals, in particular the rumen of cattle and theblind sacs (ceca) of termites, wood roaches, rabbits, ratsand pigs (Latham, 1979). The fate of surfactants in these en-vironments is beyond the scope of this paper.

6 Existing Methods

As with aerobic biodegradation test systems, a hierarchy ofscreening, inherent and simulation tests has been proposedfor anaerobic conditions. Screening tests are characterisedprimarily by a high test substance to biomass ratio, while in-herent and simulation tests aim to reach realistic concentra-tion ranges of the chemical and the bacterial biomass. In ad-dition, screening and inherent tests usually have a relativelysimple test design (e.g. batch tests, which are not capable torepresent all the dynamic processes of anaerobic degrada-tion) making them suitable for routine testing, whereas si-mulation tests necessitate the use of 14C labelled materialsor specific analytical methods.

6.1 Screening tests

The available anaerobic biodegradation screening tests arebased on the determination of the final gaseous products ofthe anaerobic degradation process under methanogenic con-ditions, i. e. carbon dioxide (CO2) and methane (CH4).However, the test conditions differ considerably from the si-tuation in a real digester. The diluted sludge inoculum cor-responds to about 10% or less of the real digester sludgeconcentration. In addition, for analytical reasons, the testcompound concentration is usually in the range of 20–100 mg of organic carbon/l, and is significantly higher thanthe concentrations usually found in digesters. Therefore, insome cases inhibitory effects are to be expected and alsohave been observed in these screening tests. This has to betaken into account when negative or poor results are ob-tained in the test. From these facts it is understandable thatanaerobic screening tests are more stringent than test sys-tems simulating realistic environmental conditions. As inaerobic screening tests, it can be concluded that a positivetest result in an anaerobic screening test is indicative foranaerobic biodegradability under real environmental condi-tions whereas a negative result is not necessarily proof of re-calcitrance.



2 EEA (European Environment Agency). Urban wastewater treatment. ETC/WTRbased on Member States data reported to OECD/EUROSTAT Joint Question-naire 2002. (http://themes.eea.europa.eu/Specific_media/water/indicators/WEU16, 2004.05)

6.1.1 Anaerobic screening tests based on gas volume measurementonly

A number of anaerobic biodegradability screening test pro-cedures are based on the method described by Shelton andTiedje (1984), e.g. the test methods used in USA (ASTM,1987), EPA (1988), the UK procedure (HMSO, 1989), Bat-tersby and Wilson (1988) and the method described by Bau-mann and Schefer (1990).

Test principle

Primary anaerobic digester sludge (with 1–3% w/v dry sol-ids) is diluted to 10% with a mineral salt medium yielding asludge concentration in the test of 1–3 g of dry solids/l. Thetest vessels (serum bottles) used have a nominal volume ofabout 160 ml and contain 100 ml of the sludge mixture aswell as the test chemical at a concentration of 50 mg of or-ganic carbon/l. After sealing, the bottles are incubated atconstant temperature (35 8C) in the dark for a test periodup to 8 weeks. The gas production is measured periodicallyby determination of the pressure increase in the headspaceof the bottles using a calibrated pressure meter. The netamount of gas produced from the degradation of the testsubstance (test gas production corrected for inoculum blankgas production) is expressed as a percentage of the theoreti-cal gas production (ThGP) calculated from the chemical for-mula of the substance taking into account the theoreticalratio of CO2 and CH4 formed in the digesting process (Bus-well equation) and the empirical solubilities of the gases inthe test medium.

Buswell equation:

CnHaOb + (n – a/4 – b/2) H2O

fi (n/2 – a/8 + b/4) CO2 + (n/2 + a/8 – b/4)

Technical aspects

The method is applicable to water soluble and poorly solublesubstances provided the concentration of the test material(50 mg of carbon/l) is not inhibitory to the anaerobic organ-isms. Knowledge of the chemical structure or, at least, of theempirical formula of the test compound is necessary so thatthe theoretical maximum amount of the gaseous final pro-ducts (CO2 and CH4) can be predicted employing the Bus-well equation. Since the test procedure only measures theheadspace gas pressure and volume, respectively, it is neces-sary to make assumptions about the relative solubility of thegases in the test mixture.

The advantage of relatively simple technical require-ments for conducting the test faces a few drawbacks:

i) If the chemical formula of the test material is unknown,the Buswell equation cannot be applied and, thus, theamounts of the individual gases cannot be calculated.

ii) The ratio of CO2 and CH4 evolved in the test may differmore or less considerably from the Buswell equation aswas shown in experiments with radio labelled test sub-stances (ECETOC 1988). If the ratio cannot be predictedreliably, the amounts of the two gases in solution cannotbe calculated.

iii) The true amount of dissolved CO2 may be variable sincethe solubility of this gas depends on a number of factors(e.g., temperature, pressure, pH, ratio of headspace/liq-uid volume, thermodynamic equilibrium established be-tween CO2 and carbonates/bicarbonates of Ca and Mg)not measured in the test. For that reason, the prescribedtest conditions have to be rigidly adhered to (Shelton andTiedje, 1984; Battersby and Wilson, 1988).

Nevertheless, in spite of these problems the method iswidely applicable and considered to be a reasonably accuratescreening procedure for the evaluation of the anaerobic de-gradation of test materials.

Test modifications

The method described by Baumann & Schefer (1990) differssomewhat from the previously described procedure by usingan extended Buswell equation to calculate directly the CH4

production from the chemical oxygen demand of the sub-stance or product tested. In addition, a lower inoculum con-centration (0.5 g/l) is used and the problem of the solubilityof CO2 is overcome by adding NaOH to the digesting mix-ture when gas production has reached a plateau. The testflasks, fitted with stirrers, are larger than in the afore-men-tioned method (250 ml nominal volume containing 200 mlof liquid) and the gas production is measured by means ofa mercury manometer fitted to each flask.

6.1.2 Anaerobic screening tests based on gas production measurementin the gas and the liquid phase

The basic pertinent test system is the so-called ECETOC(1988) test developed and published by the European Centrefor Ecotoxicology and Toxicology of Chemicals. The methodhas been ring-tested and is standardised as ISO 11734method (ISO 1995)”. The OECD 311 test guideline (2003)is an adaptation of ISO 11734. The principle on which theECETOC test and the mentioned standard test methods arebased upon is shown in Figure 4.

ISO/DIS 14853 (1999) represents another draft standardwhich is essentially identical with ISO 11734 but offers, inaddition to the manometric method, a volumetric measure-ment of the biogas production.

Test principle

A known volume of washed anaerobic sludge (1–3 g/l totalsolids), suspended in an oxygen-free mineral medium, isplaced in a suitable vessel (nominal volume 0.1–1 l) leavinga headspace (10–40% of the volume of the vessel) intowhich any gases produced can evolve. Prior to sealing, asmall amount of the test compound is added to give a con-centration of 20–100 mg of organic carbon/l. Controls(without test compound) are prepared in the same manner.Usually the test is carried out using several (e.g. 5) replicateseach of controls and test assays. The vessels are incubated atconstant temperature (35 € 2 8C) normally for periods of upto 60 days. The headspace pressure resulting from the pro-duction of gas is measured with a pressure transducer eitheronce a week or at the end of test. In addition, the DIC (dis-solved inorganic carbon) content of the digester liquid is de-termined at the end of the test. From the cumulative net gasproduction (net gas = gas in the test vessels minus gas incontrols) it is possible, by application of the gas laws, to cal-culate the amount of test compound-derived organic carbontransformed to gaseous one-carbon products (CH4, CO2). Ingeneral, the incubation is finished when the cumulative netgas production curve shows a plateau (i. e. when the gas pro-duction rate in controls and test vessels is virtually compar-able. The net DIC formation is obtained as the difference ofDIC concentrations between the test and control vessels.The extent of anaerobic ultimate degradation is calculatedby comparison of the amount of carbon equivalent to netgas and DIC production with the initially added organic car-bon content of the test chemical.



Technical aspects

The method is applicable to water soluble and poorly solublesubstances provided the test concentration (20–100 mg C/l)is not inhibitory to the test inoculum. As with the previoustest, inhibition can be easily recognised by the observation ofa negative net gas production, i. e. a gas production which ishigher in the control vessels than in the vessels containingthe test substance.

The organic carbon content of the test substance must beknown; additional substance information (chemical compo-sition, theoretical CO2/CH4-ratio according to the Buswellequation) is not necessary. This is a considerable advantageover anaerobic screening methods only based on gas volumemeasurement (cf. 6.1.1) since neither correction factors forCO2 solubility nor assumptions of the theoretical ratios ofCO2 and CH4 formation are necessary. Thus, the ECETOC/ISO 11734 method is recommended when more accuratevalues are needed for chemicals having a known empiricalformula, or if a value is required for a test material of un-known composition but its percentage carbon is known.(Painter 1994).

The ECETOC test and the standard tests based uponhave been adopted by many laboratories in Europe for sev-eral years so that its applicability and practical usefulnesscan be considered as broadly approved. Information aboutthe reproducibility of the test can be obtained from the de-gradation data reporting on, respectively, the results of inde-pendent tests of the same substance and the mean degrada-tion value including its standard deviation when multipleparallel assays are made (Pagga and Beimborn, 1993). Ac-cording to these data it can be concluded that the ECETOC/ISO 11734 test provides reproducible results showing a var-iation which is typical of biological test systems.

The ECETOC/ISO 11734 test is, like other biodegradationscreening tests, a relatively simple test system not requiringhighly specialised technical staff, but nevertheless it does re-quire sound expertise. Thus, it is strongly recommendedthat inexperienced staff is trained in the use of the test bytesting well investigated model substances and comparingthe obtained results with published data.

Test modifications

A modified ECETOC test method was used by Madsen et al.(1995) to determine anaerobic biodegradation potential indigested sludge, a freshwater swamp and marine sediment.

In the mineral medium FeCl2 was replaced by a trace ele-ments mixture. The test medium was inoculated with 10–20% fresh or washed domestic digester sludge (1.5 or0.15 g of dry solids/l), with 5% of the freshwater swamp(24–88 g of organic carbon/kg), or with 10% of the marinesediment (9 g of organic carbon/kg), respectively. The gasproduction in the headspace was measured by a pressuretransducer while CH4 was determined at the termination ofthe incubation period by gas chromatography. Dissolved in-organic carbon was quantified at the end of the test by acid-ification of the liquid and subsequent gas pressure measure-ment.

The authors suggest that the digester sludge inoculumconcentration be reduced to 0.15 g of suspended solids/l inorder to eliminate the sludge washing step usually requiredin order to reduce the amount of inorganic carbon when ahigher concentrated inoculum is used. This however in-creases the test concentration to biomass ratio unless thetest concentration is lowered accordingly.

6.1.3 Predictive value of anaerobic biodegradation screening test data

As already mentioned a common feature of biodegradationscreening tests is that they do not reproduce realistic envir-onmental conditions and are more stringent than the testssimulating the real world. Therefore, a poor degradation re-sult is not necessarily a proof of recalcitrance in the real en-vironment. Examples of anaerobic biodegradation resultssupporting this argumentation have been published (Birchet al. 1989, Steber 1991, Steber and Birch 1995a). On theother hand, a positive result in an anaerobic screening testcan be considered as highly predictive for extensive biode-gradation in anaerobic environments.

Although the production of CH4 and CO2 is the evalua-tion criterion in the anaerobic screening tests, positive re-sults in such tests may have relevance beyond methanogenicsystems. The major part of the anaerobic biodegradationroute (hydrolysis, fermentation, and acetogenesis) is com-mon to all anaerobic pathways (cf. Fig. 2). These only differbasically in terms of the final stage of the process, i. e. by thedifferent utilisation of the final electron acceptors like car-bon dioxide for methane production, sulfate for sulphideformation and nitrate for the production of ammonia andmolecular nitrogen, respectively. Thus, it can be expectedthat a chemical being degraded in the discussed anaerobicscreening tests will also undergo biodegradation in those en-vironmental conditions where the final degradation isbrought about by denitrifying or sulfate-reducing bacteria(Steber et al. 1995b).

An additional aspect of the environmental relevance ofanaerobic screening test results also has to be kept in view.The discussed tests determine the ultimate biodegradationof a chemical by measurement of the production of the finalgaseous products, i. e. methane and carbon dioxide. There-fore, it has to be acknowledged that the bacterial transforma-tion of the parent chemical (primary biodegradation) byanaerobic bacteria is not reflected by these data. Thus, evena poor degradation result in such a screening test does notnecessarily indicate for anaerobic persistence of the parentcompound. In other words, if the ultimate degradation of asubstance under these test conditions is significant (‡20%gas production) it can be concluded that the primary biode-gradation of this chemical is virtually complete as at least20% of the carbon atoms of each molecule have been miner-alised. Obviously, this assumption is only valid for purecompounds.



Figure 4 Principle and design of ECETOC Test/ISO 11734

6.2 Simulation tests

6.2.1 Introduction to anaerobic simulation tests

The motivation to move beyond the screening level to highertesting tiers can be:y The need for increased environmental realism – includ-

ing obtaining realistic kinetic information for differentanaerobic environments, inocula, reactor types or testconditions,

y To avoid inhibition of the anaerobic microorganisms bythe test material by working at a lower and more realistictest material to biomass ratio,

y To increase the chances of acclimation by exposing abroader range or organisms

y To improve the signal/noise ratio of the test,y To study the CO2 /CH4 ratioy To study the kinetics of the process.

The following test systems will be discussed:

1) 14C-Anaerobic Digester Simulation Test (batch of fed-batchsystem)

2) Anaerobic Transformation test in Aquatic Sediment Systems(OECD Guideline 308)

3) Continuously operated anaerobic reactors4) Biological nutrient removal Simulation Tests and other systems

Only the Aquatic Sediment Test (OECD308) has been ac-cepted for international standardisation. Some of the othersystems are under evaluation as potential standard methods.

6.2.2 Test systems

6.2.2.1 14C-Anaerobic digester simulation test (see Figure 5)

This test system also known as Anaerobic MineralizationTest, assesses the mineralization of 14C-radiolabelled testchemicals to CO2 and CH4 under anaerobic (methanogenic)conditions. (Steber & Wierich, 1987; Gledhill, 1995; Nuck &Federle, 1996). The inoculum is usually obtained from anactive digester but in principle the system can also be usedto simulate other anaerobic compartments, such as septictanks or sediments. The system follows the formation ofCO2 and CH4 separately over time. The underlying principle

of this method is that the headspace of the anaerobic vesselis continuously purged with N2, which is passed through aseries of base traps, to capture first the 14CO2 evolved. Theeffluent gas is mixed subsequently with oxygen and passedthrough a combustion tube (CuO at 800–900 8C) to convertthe 14CH4 into

14CO2. The latter is trapped in a second ser-ies of base traps (see Figure 5). The test can be run in abatch, or fed-batch mode (with addition of fresh sludge/testmaterial at regular intervals). The technical details of thissystem are provided in Steber and Wierich (1987) and Nuckand Federle (1996).

The anaerobic digester system, if well constructed andoperated, provides excellent recovery of the radioactivegases, and therefore good mass balances. Given the simula-tion of in-situ conditions and the realistic ratio of test sub-stance: biomass, the system can generate relevant kineticdata (Figure 6).

6.2.2.2 Anaerobic transformation test in aquatic sediment systems(OECD guideline 308)

This guideline describes a laboratory test method to assessaerobic and anaerobic transformation of organic chemicals



Figure 5 The anaerobic digester simulationtest

Alkyl ethoxy sulfate (AES) – % mineralized in anaerobic sludge(Nuck & Federle, 1996)

Figure 6 Example of the anaerobic mineralization of A24E3S (ethoxylate la-belled) to CO2 and CH4. Results are presented as % of the initial counts added

in aquatic sediment systems. An OECD Workshop on Soil/Sediment Selection, held in 1995 agreed, in particular, onthe number and type of sediments for use in this test.

The surface layer of sediment can be either aerobic oranaerobic, whereas the deeper sediment is usually anaero-bic. The aerobic test simulates an aerobic water column overan aerobic sediment layer that is underlain with an anaero-bic gradient. The anaerobic test simulates a completelyanaerobic water-sediment system.

Principle of the test:The method described in this guideline employs an aerobicand an anaerobic aquatic sediment system which allows:

(i) the measurement of the rate of transformation of thetest substance in a water-sediment system,

(ii) the measurement of the rate of mineralization of thetest substance and/or its transformation products (when14C-labelled test substance is used),

(iii) the identification and quantification of transformationproducts in water and sediment phases including massbalance (when labelled test substance is used)

(iv) The measurement of the distribution of the test sub-stance and its transformation products between the twophases during a period of incubation in the dark (toavoid, for example, algal blooms) at constant tempera-ture.

The study is typically conducted in a biometer appara-tus, or in gas flow-through systems. Half-lives, DT50, DT75

and DT90 values are determined where the data warrant,but should not be extrapolated far past the experimentalperiod.

To the best of our knowledge and at the time of writingthere was no OECD 308 study available on surfactants, butthe system potentially lends itself to anaerobic sediment de-gradation studies for this class of chemicals.

6.2.2.3 Continuously operated anaerobic reactors

Several types of semi-continuously and continuously oper-ated anaerobic reactors have been used in the literature toobtain information on the anaerobic degradability of surfac-tants.

Continuously operated fixed bed systems

Continuously operated fixed bed systems have been pro-posed by e.g. Bouwer and McCarthy (1983), Wagener andSchink (1987) and Baumann and Mueller (1997) to studythe anaerobic degradation of soluble chemicals or waste-waters. The reactors were filled with sinter-glass rings orany other suitable support material. Inoculation was with di-gester sludge and the system is maintained on a syntheticmedium. In the Wagener and Schink approach, the totalbiogas volume was monitored, while in the EMPA system(Baumann and Mueller, 1997) the biogas was collected in aglass tube containing sodium hydroxide to absorb CO2. Theanaerobic degradability of a test substance can be calculatedfrom the maximum theoretical volume of biogas/CH4 andthe amount of biogas/CH4 actually produced.

The continuous anaerobic fixed bed system can be usedto perform simple mass balance studies, or according to theauthors, to gain information on hydrophilic and hydropho-bic metabolites (provided adequate analytical methods areavailable). A typical test concentration for a non-labelled(non-toxic) chemical is in the order of 250–500 mg/l asCOD.

Continuous stirred tank systems

Metzner & Lemmer (1997) illustrate a dynamic semi-conti-nous system loaded intermittent with sludge from munici-pal sewage digester. Due to varying sludge qualities of dif-ferent origin, semi-continous systems are difficult tostadardize. Also high background levels of organics, e.g. sur-factants cannot be avoided. For this reason Metzner (2001)suggested to use synthetic sludge in a thermostatically con-trolled continous system. The possibility of automatizationleads to a lower susceptibility to faults and the advantage ofbetter standardization. Austermann-Haun et al. (2004) re-port on the use of a thermostatically controlled continuouslystirred reactor system where the amount of biogass formedis measured. No specific results for surfactants were re-ported. This system can be run continously and/or semi-continously as a degradation and inhibition test.

Metzner (1998) describes the absence of degradation andthe inhibition at higher concentrations of LAS in an anaero-bic, semi-continously fed stirred tank reactor.

Between 2000 and 2003, three separate papers were pub-lished by a group in Denmark describing the disappearanceof LAS in laboratory continuous stirred tank reactors (Angel-idaki et al., 2000; Hagensen et al., 2002; Mogensen et al.,2003). The experiments consisted of 3.5 L reactors semi-con-tinuously fed with sludge solids spiked with LAS and moni-toring by HPLC the level of LAS in the effluent over time.The reactors had hydraulic retention times of 15 days andwere operated at 37 8C. They aimed to simulate a singlestage anaerobic digestor common in treatment plants. Onlya limited removal of LAS was observed in this reactor sys-tem. The exact mechanism of removal (degradation and/orsorption) was not fully clarified.

Upflow anaerobic sludge blanket systems

The same research group from Denmark published two se-parate papers on the biodegradation of LAS in UASB reac-tors (Mogensen & Ahring, 2002; Mogensen et al. 2003).UASB reactors are commonly used to treat industrial waste-waters and less for domestic wastewater. They containgranulated sludge held in suspension by the upflowingwater. The authors chose this type of reactor to maximizethe bioavailability of LAS. Removal was determined in thesterile and bioactive reactors. The effluents levels from thebioactive unit levelled off at 6 mg/l, while those from thesterile unit continued to increase up to 9 mg/l at 312 hours,which suggests some degradation of LAS. Analyses of theeffluent from the bioactive unit revealed a transient 24% in-crease in volatile fatty acid production relative to the sterilecontrol upon addition of LAS to the feed. In addition, totalion chromatograms revealed several peaks in the effluent ofthe bioactive unit not found in the effluent of the sterile con-trol or a unit fed only minimal salts medium without LAS.One of the papers identifies the presence of benzene sulfo-nate (Mogensen & Ahring 2002) and another identifies thepresence of benzaldehyde (Mogensen et al. 2003). Neitherpaper mentions both materials together. There is no indica-tion on the level of either metabolite that can provide insightinto the stoichiometry of the reactions.

A paper by a group in Spain (Sanz et al, 2003) also refersto UASB reactors, operated at 30 C. One reactor was fed or-ganic nutrients plus 4–5 mg LAS/l and the other was fedonly LAS in mineral medium. Based upon the total massbalance, 64% of the LAS was lost from the first reactor and85% from the LAS-only reactor. There was no sterile control,nor data on gas production and metabolites in the effluent.



6.2.2.4 Biological nutrient removal STP simulation tests:Behr and CAS-UCT units

Wastewater treatment systems with biological nutrient re-moval are becoming quite common across Europe (http://epp.eurostat.ec.europa.eu). The Behr- and CAS-UCT testsystems are modifications of the OECD303A method and al-low to simulate the behaviour of chemicals (e. g. surfactants)in laboratory-scale sewage treatment plants with anoxic(Behr unit) and anaerobic plus low dissolved O2 zones(CAS-UCT system) (Schowanek et al., 1996). These are typi-cal configurations for single-sludge wastewater treatmentplants with biological N/P removal. While these are not en-tirely anaerobic biodegradation test systems, they have beenincluded in this chapter for information purposes. TheBehr- and CAS-UCT laboratory units can also be constructedin a way which allows 14C-mass balance studies.

6.2.2.5 Others systems

Various other anaerobic reactor designs (e.g. septic tank,anaerobic contact process, aquifer column,) can also be oper-ated on a laboratory-scale to study the behaviour of chemi-cals in these specific systems or environments (e.g. Kuhnet al., 1988; Heijnen et al., 1991).

Since these would be considered as non-routine tests,they will not be covered further.

7 Interpretation of Available Data on AnaerobicBiodegradation of Surfactants

This chapter gives a compilation of existing literature dataon anaerobic biodegradation of all surfactant classes.

Data on laboratory tests are shown in individual tablestogether with a discussion of the results.

Monitoring data of surfactants in anaerobic compart-ments reported in the literature do not necessarily reflectthe biodegradation behaviour of surfactants under strictlyanaerobic conditions. For example, temporary aerobic condi-tions in an otherwise anaerobic environment may influencethe results towards a high degree of degradation. Thereforesuch monitoring data (see Appendix) may not be used as aproof of biodegradability/recalcitrance in anaerobic environ-ments, but may form an additional support of the conclu-sions drawn from the laboratory test results.

7.1 Anionics

7.1.1 Sulfonates

In tables: Separate lab studies from the field studies

Sur-fac-tanttype

Characteri-sation

Test type Test conc. Inoculum(dw) conc.

TestDura-tion

Temp. Results Remarks References

Screen-ing

Simulation in mg/lactivematter

mg/lcar-bon

in g/l Days 8C

LAS C107C13commercial

ECE-TOC

50 1–5 49 35 0 Steber (1991)

LAS ring-14C digester 10 20 27 37 0.3 14C LAS Steber (1991)

LAS ring-14C field system 0.5 WW pondsed.

87 22 0 14C LAS Federle &Schwab (1992)

LAS C12 LAS ECE-TOC

50 60 inhibition bio-gas formation

Battersby &Wilson (1989)

LAS C8 LAS ECE-TOC

50 60 inhibition bio-gas formation

Battersby &Wilson (1989)


field system septic tank 97 (a) 10% 14C LAS Klein andMehaughey

(1964)


field system 10 99.7 (a) Larson et al.(1989)

LAS C12 LAS researchstudy

5 sedimentslurry

6 100 (k < 0.23 d-1)(b)

14C LAS Heinze &Britton (1994)


researchstudy

5–20 0.02–0.1 29 h 25 100 (k = 0.14 h-1)(c)

specific LASand S

Denger & Cook(1999)

LAS 2-Ph C10LAS, C12and C14

Birchet al.1989

researchstudy

20 Anaerobicsludge

49days

Negligible both forLAS and for SPCSame results withPh in position 5

14C LAS en-hances biogasproduction

Garcia M. T.et al. (2005)

LAS C12 LAS a-Semi-con-tinuous stir-red tak reac-tor CSTRb-UASB

100

10–20

Anaerobicsludge

15days

12hrs

37

55

20% biodeg.

37% biodeg. UASB

Mogensen A. S.et al. (2003)


UASB upflow anaero-bic sludge

bed

1 UASB ther-mophilicgranularsludge

12hrs

40% biodegrada-tion

Benzene sul-fonic acid andbenzaldehydewere found

Mogensen A. S.and Ahring B. K.

(2002)



Sur-fac-tanttype

Characteri-sation

Test type Test conc. Inoculum(dw) conc.

TestDura-tion


Screen-ing

Simulation in mg/lactivematter

mg/lcar-bon

in g/l Days 8C

LAS LAS C12LAS 14C12

CSTRUASB

20 Digestedsludge

8 hrs 37and55

40–80% removal50% absorbed on

sludge5.6% degradation

by 14CO2

Specific LASanalyses

Angelidaki, I.et al. (2003)and (2004)

LAS LAS C12 CSTR Anaerobicsludge

14–25% primarybiodegradation

Specific LASanalyses

Haggensen, F.et al. (2002)


ECE-TOC

Anaerobicdigester

Poor primary bio-deg. for single LAS

isomers

Garcia, et al.(2004)

LAS LAS C10–13

Laboratorystudy

Marine se-diments

LAS absorption onmarine sedum. and

organisms

Equilibrium:12 hrs for LASand 18 hrs for

SPC

Saes M. et al.(2003)

LAS LAS C10–13

ECE-TOC

Anaerobicsludge

4–22% primary 4–26% inhi-bition of biogasproduction

Garcia et al.(2006)

LAS LAS C10–13

ECE-TOC

Laboratorystudy on LASeffect on theanaerobic di-gestion ofsewagesludge

Anaerobicsludge

LAS at conc. < 5–10 g/kg dw in-

crease biogas pro-duction; inhibitionat higher concen-

trations14.5 m/l estimated

LAS toxicity onanaerobic micro-

orgs

Garcia, Camposet al. (2006)

LAS LAS C10–13

commercial

Laboratorystudy

10–50 Marine se-diments

165 79 degradationwithin 165 days viathe generation ofsulfophenyl car-boxylic aids

Proof of theprimary biode-gradation ofLAS under

anaerobic con-dit.

Pablo A. Lara-Martín et al.(2007)

MES 2-sulpho(14C) palmi-tic Me ester

digester 10 20 28 35 0.6 14C MES Steber &Wierich (1989)

SAS C147C17 ECE-TOC

17 0 Bruce et al.(1966)

a) Oxygen – limited conditionsb) Mineralization kinetic study in different oxygen-limited conditionsc) Desulfonation kinetic study using an enriched and isolated strain. Primary biodegradation

Conclusions on the anaerobic biodegradability of sulfonates

1) Laboratory data

Sulfonates are not degraded significantly under the anaero-bic conditions of the laboratory test methods or in anaerobicdigesters of sewage sludge (Steber, 1991, Mogensen 2002,Mogensen 2003, Angelidaky 2004, Garcia 2004 and Garcia2005).

In the real environment and also in field system tests,however, oxygen-limited conditions may be more commonthan rigorously anaerobic conditions. In such conditionssulfonates mineralize even if the rate is not as rapid as thatobserved under aerobic conditions. Once sulfonate biodegra-dation has been initiated in aerobic or oxygen-limited condi-tions, the intermediates can continue to biodegrade anaero-bically. This is the reason why in some simulation teststhese chemicals can show mineralization results, if someoxygen diffusion had occurred or if limited-oxygen condi-tions had been created (Larson, 1989; Heinze and Britton,

1994). On the other hand, it must be also considered thatonly 0.2 mg/l of dissolved oxygen can support biodegrada-tion if O2 recharge to the aquifers is not limited (Salanitro1997).

It was observed that LAS can inhibit biogas formation inanaerobic biodegradation batch tests if present at high con-centrations (Battersby and Wilson, 1989). It has also beenfound however that LAS can increase the biogas productionwhen present at lower concentrations. In fact, Garcia andCampos 2006 determined that the inhibition starts at LASconcentrations of about 5–10 g/kg dw of LAS as Na saltswhile, at lower concentrations the biogas production is in-creased. Therefore, at the usual LAS concentration range insewage sludge, no adverse effects on the anaerobic digestersof a wastewater treatment plant (WWTP) can be expected. Inaddition, in real anaerobic environmental compartments(e.g. STP anaerobic digester), however, sulfonates haveshown not to inhibit biogas formation even at high concen-tration (>30 g/kg dw) because they are present as Ca saltswhich are poorly soluble and less bio-available (see 8.2).



It has also been demonstrated that not all desulfonationreactions require oxygen. Desulfonation of alkylsulfonatesas well as of LAS has been reported to occur by anaerobicbacteria in the laboratory (Denger and Cook, 1999). Thereis recent evidence that anaerobic desulfonation can takeplace. Desulfonation with assimilation of the sulphur moi-ety by strictly anaerobic bacteria (Chien et al., 1995; Dengeret al., 1996; Denger and Cook, 1997) was followed by the re-duction of the sulfonate group as a source of electrons andcarbon under anaerobic nitrate-respiring conditions (Lie etal., 1996; Laue et al., 1997; Denger et al., 1997). There is noevidence however that such mechanisms would also occur atsignificant rates under real-world conditions.

Two breaking through publications of Pablo A. Lara-Mar-t�n et al. (2007) proved for the first time the degradation oflinear alkylbenzene sulfonates (LAS) under anaerobic condi-tions through the generation of sulfophenyl carboxylic acids(SPCs), together with the presence of metabolites and theidentification of microorganisms involved in this process.Laboratory experiments performed with anoxic marine sedi-ments spiked with 10–50 ppm of LAS demonstrated that itsdegradation reached 79% in 165 days via the generation ofSPCs. Almost all of the added LAS (>99%) was found to beattached to the sediment while the less hydrophobic SPCswere predominant in solution, as their concentration in-creased progressively up to 3 ppm during the full course ofthe experiment. The average half-life for LAS was estimatedto be 90 days. The researchers found that the degradation isreduced when the amount of LAS in sediment exceeds20 ppm, as this concentration could be considered toxic tocertain marine microorganisms. However, these LAS valuesare unlikely to occur in aquatic ecosystems except for specif-ic locations near untreated wastewater discharges. Sulfate-reducing and methanogenic activities proved to be intenseduring the experiment. Several sulfate-reducing bacteriaand firmicutes/clostridia were identified as possible candi-dates for effecting this degradation.

2) Monitoring data (see Appendix)

SludgeLAS concentrations in sludge have been reported in severalsituations ranging from < 100 mg/kg to a maximum of30,000 mg/kg (Berna, 1989) depending on emission vol-ume, STP operating conditions and water hardness. Aerobicstabilised sludge always has a LAS content typically lowerthan 500 mg/kg dry matter, whereas anaerobically digestedsludge has a LAS concentration in the range 1,000–30,000 mg/kg, with a mean value around 5,000 mg/kg.(Waters, 1995 and McAvoy et al., 1993, and Schowanek etal. in press). The latter authors present a distribution ofLAS concentrations found in anaerobic sludges, based ondata reported in the scientific literature and (industry) re-ports, covering different European countries, and the timeperiod 1988–2006. Results are overall quite consistent be-tween different studies, and between the EU and the US.

A few studies have attempted to make a LAS mass bal-ance over full scale anaerobic digesters. A low degree of re-moval is measured, ranging from 0–35% (Berna, 1989 andGiger, 1987). These data fit broadly with the results from la-boratory screening or digester tests, and biochemical in-sight, which indicate that LAS, as well as any other sulfonatesurfactants, does not degrade under strictly anaerobic condi-tions. It is not entirely clear to which process (as) the smalldegree of removal observed in field systems can be adscribed(binding, humification, co-metabolism, and anaerobic desul-fonation). LAS concentrations on a wet sludge basis actually

increase versus fresh aerobic sludge in an anaerobic digesterdue to the dewatering and digestion of solids.

We are aware of only another study on anaerobic degra-dation of sulfonates in full scale systems, not related to LASbut instead to SAS (Field, 1992). The SAS monitoring inanaerobic sludge ranges from 270 to 800 mg/kg. Like LAS,there is no evidence for a significant anaerobic degradationof SAS.

SedimentsLAS has been found in river sediments. The levels foundjust below an STP outfall range from zero to a maximumvalue of 174 mg/kg, (Rapaport, 1990). Downstream thesame STP outfall the LAS concentrations in sediments dropto 5–11 mg/kg (Rapaport, 1990). In Po river sediments, con-centrations from 0.4 to 4.7 mg LAS/kg were detected (Cavalli2000). No concentration effect was observed along 500 km ofthe river, indicating that there is no LAS accumulation de-spite the high load from this big catchment.

LAS has also been detected in marine sediments at con-centrations ranging from not detectable to few ppm, (DelValls 2002, Folke 2003, Bester 2001, Leon 2001, Lara-Matin2006). Higher concentrations were found only in polluted lo-cations like 10 mg/kg in one site used for direct discharge(Lara-Martin 2005) or 20 mg/kg in a Danish port (Folke2003).

No mass balance studies are known to us. Several inter-esting studies have looked at the levels of LAS and other sur-factants in cores or river/lake sediments as accumulatedover time. Residual LAS levels can be measured, and in thetrends over time (i. e. depth) clear records can be seen re-flecting the switch from branched alkyl sulfonates to LAS(Schçberl, 1996; Reiser, 1995).

SoilLAS and SPC were determined in sludge amended soils of10 field sites in the North of Spain. The sludge applicationwas generally at the rate between 8 and 20 tonnes per hec-tare and the sites were treated at least once over the period1997–2002. The concentrations were in the range 0.12–2.8 mg /kg and 0.004–0.2 mg/kg. respectively for LAS andSPC. The highest concentration was found in the field thathad received sludge 10 days before the sampling (Eichhorn2005). The authors concluded that, applying the estimatedPNEC of 4.6 mg /kg (Jensen 2001) LAS do not pose a signif-icant risk to fauna, plants and essential functions of agricul-tural soils.

In another Danish study that includes 8 locations withdifferent histories, including sludge amended soil, it wasconcluded that it is not possible to note significant differ-ences between undisturbed soils used for grazing for 50–100 years and soils being moderately sludge amended. AllLAS concentrations were in the range 0.5–1.0 mg/kg. Onlyin an heavy sludge amended location was found 10–20 mgLAS/kg but it was not representative because the amend-ment of this location was carried out at a high rate/17.5 tdw/ha/y for 25 years/, therefore not respecting the currentlyadopted EU regulation (Carlsen 2002).

These low concentrations of LAS in soil can be explainedby the fact that LAS can be used by bacteria as a source ofenergy and carbon (Elsgaard 2003), and rapidly degrades insoil with a half life of ca. 10 days (primary degradation) and30 days (mineralisation) (Schowanek et al., in press).

Several risk assessments of LAS in the terrestrial envir-onment have been published over time and have introducedsubsequent refinements (Mieure et al. 1990; De Wolf andFeijtel 1998; Jensen 1999; Jensen et al. 2001; HERA 2004).



Recently, the PNECsoil has been revised upwards from 4.6to 35 mg/kg (Jensen et al., 2007). Based on this work, Scho-wanek et al. (in press) have performed a deterministic andprobabilistic risk assessment for LAS in anaerobic sewagesludge. The conclusion, backed up by relevant field evi-dence, is that LAS in anaerobic sludge does not represent asignificant ecological risk. From this perspective there is no

scientific basis for specific regulatory limits for LAS insludge as present in earlier EU draft legislation, nor theneed to consider specific sludge management options dueto the presence of LAS.

7.1.2 Sulfates

Surfactanttype

Characterisa-tion

Test type Test concentration Inocu-lumconc.

Testdura-tion


Screening Simula-tion

in mg/lactivematter

in mg/lcarbon

in g/l indays

8C %

AlcoholSulfates

C12 Digester 10 5 20 28 35 90 14C-labelled(31% CO2

+ 59% CH4)

Steber et al.(1988)

C12–13linear

C12–13 mid-chain

branched

C12–13mainly

branched inß-position

ECETOC 30 1–5 42 32.6–36.8

70

40

25

Biogas + IC Rehman et al.(2005)

C12–14 Mod.Shelton &Tiedje

96 50 1 40–50

35 77–84 CH4 Salanitro & Diaz(1995)

98–99 MBAS

C12–18 ISO14853

100 1.5 84 35 5999

Biogas + ICMBAS

Fraunhofer(2003)

C14 Digester 1 0.5 26 15 35 80 14C-labelled(CO2 + CH4)

Nuck & Federle(1996)

C14–15 Mod.Shelton &Tiedje

93 50 1 40–50

35 65–78 CH4 Salanitro & Diazl(1995)

97 MBAS

C14–1580% linear

ECETOC 30 1–5 42 32.6–36.8

60 Biogas + IC Rehman et al.(2005)

C18 ECETOC 50 29 3 56 35 88 Biogas + IC ECETOC (1988)

C18 Digester 10 6 20 28 35 94 14C-labelled(43% CO2+ 51% CH4)

Steber et al.(1988)

AlcoholEthoxySulfate

C12, xEO Mod.ECETOC

40–100 20–50

0.06–0.12

55–56

35 14–41 Biogas (in-cludes IC afteracidification)

Madsen & Ras-mussen (1994)

C12–14,2 EO

ECETOC 50 1–5 41 35 75 Biogas + IC Steber (1991)

C12–14,2 EO

ISO14853

100

50

1.5

1.5

84

119

35

35

o7360

Biogas + ICMBAS

Biogas + IC

Fraunhofer(2003)

C14–15,2 EO

Digester 1–10 0.5–5 26 17 35 88 14C-labelled(CO2 + CH4)

Nuck & Federle(1996)

C14–18,3 EO

Septictank

26–52 14–29 Nodata

240 Nodata

72–81 MBAS Painter (1992)

Conclusions on the anaerobic biodegradability of sulfates

Alcohol Sulfates

1) Laboratory data

Linear primary alcohol sulfates with alkyl chain lengths typi-cally used in detergents (C12–18) are anaerobically biode-gradable, with a conversion to CH4 and CO2 (biogas) of80–94% in tests simulating anaerobic digesters. They are

also extensively mineralised (59–88%) in screening tests,although as with other anionic surfactants, the unrealisti-cally high surfactant to biomass ratio means that low biode-gradation or inhibition of biogas production are sometimesobserved. Branching of the alkyl chain results in a reducedextent of ultimate anaerobic degradation under screeningtest conditions. The primary biodegradation of alcohol sul-fates under anaerobic conditions is very high exhibiting aMBAS reduction of ‡ 97% in screening tests.



2) Monitoring data

SludgeNo monitoring data are available.

SedimentsIn a monitoring study at the Bay of C�diz, Spain, (Lara Mar-t�n et al., 2005) AS even alkyl chain homologues from C12 toC18 AS were found in very low concentrations (<100 ppb),while no levels of AS with odd homologues were detected inmost of the sampling stations due to their low usage in com-parison to other high volume surfactants like LAS or AES.

SoilNo monitoring data are available.

Alcohol Ethoxysulfates

1) Laboratory Data

Compared to alcohol sulfates less experimental data areavailable on the anaerobic biodegradability of ethoxysulfateshaving alkyl and EO chain lengths typical of the use in deter-gents (C12–18, £4EO). However, the existing informationindicates that they are also extensively biodegraded (‡60%biogas formation) under anaerobic conditions. Low biode-gradation has been reported in some cases, e.g. for AES,but this can be attributed to the very high test substance tobiomass ratio used in the study quoted. Similarly, the re-ported data from ISO 14853 screening tests show the depen-dence of the results from the test concentration. In addition,it becomes evident that even test results with poor biogas

formation may imply a high percentage of primary biode-gradation (MBAS reduction).

2) Monitoring Data

SludgeNo monitoring data are available.

SedimentsA research group from the University of C�diz, Spain, inves-tigated the distribution of AES homologues in sedimentstaken from the coastal area of Spain (Gonz�lez Mazo et al.,2005).

Concentrations of AES in those marine sediments havebeen found to be very low in the Bay of Cadiz (Spain) as wellas in other bays of the Spanish coast. Typical concentrationvalues of AES homologues in 1 cm depth were ranging be-tween 150 and 400 ppb. The main trend along the sedimen-tary column is towards a decrease in the concentration va-lues of all these compounds with depth, with no detectionor a non possible quantification due to very low concentra-tions in most cases at 17 cm and especially 47 cm. An addi-tional monitoring study at the Bay of C�diz was carried out(Lara Mart�n et al., 2005) finding concentrations of AES of125 ppb.

SoilNo monitoring data are available.

7.1.3 Fatty acids and soaps

Surfactanttype

Characterisa-tion

Test type Test concentr. Inoculumconcentration

Testdura-tion

Temp. Re-sults

Remarks References


in mg/lactivematter

car-bon

in g/l Days 8C %

Fatty acid dodecyl (C 12) ECETOC 20 0.15 sludge 56 35 >75 gas afteracidific.

Madsen et al.(1995)

ECETOC 20 4.4 organic C/l(freshwaterswamp sedi-

ment)

56 35 >75 gas afteracidific.

Madsen et al.(1995)

ECETOC 20 0.9 organic C/l(marine sedi-

ment)

96 35 >75 gas afteracidific.

Madsen et al.(1995)

coconut(C 12–18)

ECETOC 18 0.06/0.12sludge

55 35 40–57

gas afteracidific.

Madsen &Rasmussen(1994)

(u*-14C)palmitic (C 16)

static 10 20 sludge 28 35 96.5 56.6%CH4 + 39.9%

CO2

Steber &Wierich(1987)

Soap Na-laurate(C 12)

semicont. 200 30 sludge 20 (re-tentiontime)

35 95 CH4 meas-ured, CO2

calc.

Mix-Spagl(1990)

Ca-laurate(C 12)

static 1000 30 sludge 5–6 35 90 CH4 meas-ured, CO2

calc.

Petzi (1989)

Na-palmitate(C 16)

ECETOC 70 1–5 sludge 28 35 94 gas + DIC Birch et al.(1989)

Na-palmkernel

(C 8–18)



calc.

Petzi (1989)

Na-tallow(C 16/18)



calc.

Petzi (1989)

* uniformly marked



Surfactanttype

Characterisa-tion

Test type Test concentr. Inoculumconcentration

Testdura-tion

Temp. Re-sults

Remarks References


in mg/lactivematter

car-bon

in g/l Days 8C %

Soap Na-stearate(C 18)



calc.

Mix-Spagl(1990)

Ca-stearate(C 18)

static 1000 30 sludge 10 35 85 CH4 meas-ured, CO2

calc.

Mix-Spagl(1990)

Na-oleate(C 18 unsat.)



calc.

Mix-Spagl(1990)

Na-behenate(C 22)



calc.

Mix-Spagl(1990)

Ca-behenate(C 22)

static 1000 30 sludge 10 35 90 CH4 meas-ured, CO2

calc.

Mix-Spagl(1990)

Na-Laurate(C 12)

ECETOC static 100, 200,400, 600

54 35 94–100

gas + DIC Varo et al.(2002)

Na-Laurate(C 12)

ECETOC static 1000 54 35 0 gas + DIC Varo et al.(2002)

Na-Stereate ECETOC static 100, 200,400, 600,

800

54 100 gas + DIC Varo et al.(2002)

Conclusions on the anaerobic biodegradability of fatty acids andsoaps

1) Laboratory data

Fatty acids and their sodium/calcium salts (soaps) are wellbiodegradable under anaerobic conditions. This was shownin several laboratory studies using different test methods(see Table). The data for dodecanoate and palmitate obtainedin the stringent ECETOC screening test showed very highmineralization rates (> 75% CO2 + CH4 formation) withinthe test period of 4–8 weeks. The bacterial inocula used inthese investigations were digester sludge as well as anaero-bic sediments from fresh water and marine environments.The data for C12 – 18 fatty acids reported by Madsen and Ras-mussen (1994) are lower (40–57%) but this may indicateslower biodegradation kinetics due to the unusually hightest substance/inoculum ratio used in the test. The positiveevaluation of the anaerobic biodegradability of fatty acidsand soaps was confirmed in a digester simulation studyusing radio labelled palmitate and showing an almost quan-titative ultimate degradation to carbon dioxide and methane(Steber and Wierich 1987). Additional static and semi con-tinuous digester simulation tests proved the extensive ulti-mate biodegradation of Na- and Ca-salts of fatty acids withan alkyl chain length of 8–22 carbons (Petzi 1989, Mix-Spagl 1990). While the Ca-soaps (C12, C18, C22) exhibitedhigh gas formation rates (‡ 85%) in the static system withina 10-day test duration, the semi continuously run investiga-tions with Na-soaps showed that the time needed for miner-alization was increasing with the chain length and concomi-tantly with the water solubility. It should be noted the allthese data refer to ultimate biodegradation based on themeasurement of gas production.

2) Monitoring data (see Appendix)

SludgeAnaerobic biodegradation of soap has been also observed incontinuous laboratory scale equipment as well as on com-mercial scale sewage treatment plant (D. Prats et al., 1999).Average removal in both cases was in the 64–90% rangebeing the lower values obtained for unsaturated homologuesregardless the residence time (25, 40 and 60 days). The factthat no higher removals were observed can be attributed topoor bioavailability of Ca–Mg soaps. Besides, fatty acidsmight be potentially formed during the digestion processfrom manifold chemicals of natural origin (e.g., lipids).These results confirm previous mass balances of soap basedon a monitoring study of the concentrations in sludge of adigester influent and effluent (Moreno et al. 1993).

SoilThe soap content was measured in sludge amended and nonamended soils from the southeast of Spain (Alicante). Soapconcentration in the untreated soil (blank) was 0.57 g/kg.Once the soil was amended the soap initial concentrationwas 1.47 g/kg. The concentration of soap 150 days afteramendment was below the value at the moment of theamendment and can be considered therefore as totally bio-degraded after this period. A steady state concentration ofsoap might indicate the residual soap present as non bioa-vailable form (D. Prats et al., 1999).

SedimentsThere is a paucity of data about the presence of soap in ma-rine sediments. The existing data evidence soap concentra-tios in the 500–9000 mg/kg range (Folke et al., 2003; L�pezet al., 2002). Although these studies cannot pinpoint the de-tergent-origin of the detected soap, its concentration in rela-tively uncontaminated sediments is from 15,000 up to40,000 times higher than the LAS levels.



7.2 Nonionics

Surfac-tanttype

Characterisation Test type Test Concentra-tion

InoculumConcen-tration

Testdura-tion

Temp Results Remarks References

Screening Simu-lation

in mg/lactivematter

in mg/lcarbon

days 8C

AlcoholEthoxy-lates

C 9–11, 8 EO ECETOC 20–50

0.15 g drymatter/l

56 35 >75% Inoculum:digested sewagesludge, fresh-water swamp

Madsen et al.(1995)

C 9–11, 8 EO ECETOC 20–50

0.15 g drymatter/l

56 35 30–75 Inoculum: mar-ine sediment

Madsen et al.(1995)

C 9–11, 8 EO Mod. ISO11734(1995)

20–50

0.6 g drymatter/l

55 35 65–82%biogas

Inoculum:digester sludge


C 10–12, 7,5 EO screening 10–1000

37 70% biogas Inoculum, anoxicfreshwatersediment

Wagener &Schink (1987)

C 12, 23 EO screening 10–1000

37 80% biogas Inoculum, anoxicfreshwatersediment


C 12, 23 EO simula-tion

1000 90 >90% CH4,acetate,

propionate

Laboratory-scaleanaerobic fixed-bed reactor fedwith syntheticwastewater


C 12, 9 EO 14C-screen

1,7 ug/g sedi-ment

87 22 13–40%14C; glu-cose: 40%

WW pond sedi-ment

Federle &Schwab (1992)

Isotridecanol,(5,10, 20) EO

ECETOC,modified

20 2–3 g/las solids

110 35 0–30%ThGP

Siegfried et al.(1996)

linear primaryC 12-C15 7 EO

ECETOC,modified

20 0.15 g drymatter/l

35% ThGP Inhibition duringthe first 3 weeks

Madsen et al.(1996)

Linear C 12–14,(5, 10, 20) EO

ECETOC,modified


110 35 29–94%ThGP


mono br. C 14–15, (10,20) EO

ECETOC,modified


89 35 0–23%ThGP


C 18, 7 EO 14C-sim.

10 12–25 gdry mat-ter/l

28 35 83–87%14C as bio-

gas

Steber & Wierich(1987)

Linear C 12-EOn laboratory 40 mg/l anaerobicsludge

40 85.1% Huber et al.(2000)

Alkyl-phenolEthoxy-lates

C 10–12 alkyl-phenol, 9 EO

screening 10–1000

37 45–50%CH4

Inoculum: anoxicfreshwatersediment


Nonylphenol,10 EO

ECETOC 50 1–5 mgdry mat-ter/l

84 35 20.5 ±12.6%

CO2 + CH4

Steber (1991)

Nonylphenol,9 EO

ECETOC 50 1 g drymatter/l

40–50

35 32–43%CH4

Salanitro & Diaz(1995)

Nonylphenol,0/1/2 EO

simula-tion

NP is metaboliteof APEO,

accumulates

Tschui & Brunner(1985)

GlucoseDeriva-tives

Gluco-side

Ethyl 6-O-de-canoyl glucoside

Mod. ISO11734(1995)

20 56 35 59–65%biogas

Inoculum: diges-ter sludge


APG (branched)C8, DP = 1.6

ECETOC,modified

30–40


22% ThGP Primary domesticsludge

Madsen et al.(1996)

APG (linear)C 12–14,DP = 1.4

ECETOC,modified

30–40



Madsen et al.(1996)



Surfac-tanttype

Characterisation Test type Test Concentra-tion

InoculumConcen-tration

Testdura-tion

Temp Results Remarks References

Screening Simu-lation

in mg/lactivematter

in mg/lcarbon

days 8C

Gluco-side

C 12 Ethylgluco-side monoester

EGE

ECETOC,modified

30–40



Madsen et al.(1996)

C 12 6-O-Ethyl-glucosidemonoester

ECETOC 20–50

0.15 g drymatter/l


Madsen et al.(1995)

C 12-C14 APG ECETOC 20–50

0.15 g drymatter/l


Madsen et al.(1995)

C 10 6-O-Ethyl-glucosidemonoester

ECETOC 20–50

0.15 g drymatter/l

55 35 >75% Inoculum:digested sewagesludge, marinesediment, fresh-water swamp

Madsen et al.(1995)

C 12–14-APG ECETOC 100 3 g drymatter/l

56 35 84 ± 15%(CO2) +(CH4)

Steber et al.c. (1995)

C 8–10-APG ECETOC 100 3 g drymatter/l

56 35 95 ± 22%(CO2) +(CH4)

Steber et al.c. (1995)

Glu-coseAmide

C 12 GlucoseAmide

14C-sim.

1 35 86 ± 0,3%(CO2 +CH4)

Federle and Nuck(1997)

AmineOxides

Dimethyl Dode-cyl Amine Oxide

14CDiges-ter

1 ~20 7 35 >80% Vandepitte andDebaere (1997)

Alkyl-ethanol-amides

Coco monoetha-nolamide

ISO 11734 20 1–3 56 35 81% of the-oretical gasproduction

Madsen (2001)

DP = degree of polymerizationAPG = alkylpolyglucosideTHGP = theoretical gas production

Conclusions on the anaerobic biodegradation of nonionicsurfactants

Alcohol Ethoxylates

1) Laboratory data

Alcohol ethoxylates (linear C9–C18-alcohols, 5–23 EO) arewell biodegradable in anaerobic screening tests. Degradationof usually >70% (biogas) has been reported in digested sew-age sludge and freshwater sediment.

Degradation >80% (biogas formation and 14C) in diges-ter simulation tests have been reported in the case of lauryl-alcohol ethoxylates and stearylalcohol ethoxylates. In a studyon the biodegradation mechanisms of linear alcohol ethoxy-lates under anaerobic condition it was concluded that thefirst step of anaerobic microbial attack is the cleavage of theterminal EO unit, releasing acetaldehyde stepwise, andshortening the ethoxy chain until the lipophilic moiety isreached (Huber et al., 2000).

2) Monitoring data

SludgeActual data on the concentration of alcohol ethoxylates in di-gester sludge support the conclusion that alcohol ethoxylatesare well biodegradable in full scale digestors (Klotz, 1998).

In a number of EU countries concentrations of alcoholethoxylates were monitored in inlet and outlet sludgesstreams. The removal of alcohol ethoxylates by anaerobic di-gestion was on average 82% (range 61–93%). The appliedanalytical method determined alcohol ethoxylate componentswith C12–C18 alkyl chain lengths and an ethoxylate chainlength of 4–20 EO in sludge samples (Matthijs et al., 2004).

Concentrations of alcohol ethoxylates were monitored ininlet and outlet sludges streams of a number of STP in dif-ferent EU countries. The removal of alcohol ethoxylates byanaerobic digestion was on average 82% (range 61–93%),(Matthijs et al., 2004). Removal of alcohol ethoxylates inanaerobic sludge of a single STP was also determined inthe range of 54–74% by Bruno et al. 2002. Alcohol ethoxy-lates were also detected by Cantero et al. in anaerobic sludgeat concentrations of 23–141 mg/kg while nonyl and octylphenol ethoxylates were in the range 11–151 and 100–138,respectively. The removal of alcohol ethoxylates’ homologueswas >99% in nine US sewage treatment withactivatedsludge process. The total average concentration of alcoholethoxylates in effluent was 0.9 lg/l compared to 15.6 lg/lin trikling filter (Morral et al., 2006). A monitoring of envir-onmental fingerprints of alcohol ethoxylates in Europe andCanada was published by Eadsforth et al. 2006.

No systematically investigations into the influence of thebranching degree of the alcohol moiety on the anaerobic bio-



degradability of alcohol ethoxylates are available. M�ller(Thesis, 2000) published some results of linear, single- andmultiple-branched alcohol ethoxylates using sludge fromdifferent digester plants. The ethoxylation degree rangedfrom 5 to 20 mol EO, which did not influence significantthe biodegradability. High degradation rates were found forlinear alcohol ethoxylates. Overall, the anaerobic biodegrad-ability of alcohol ethoxylates decreased with increasingbranching degree of the alcohol moiety from linear to multi-ple-branched alcohols.

SoilThere are no published relevant data of alcohol ethoxylatesin soil. Alcohol ethoxylates were detected at 194 ng/l in soilinterstitial water and 710 ng/l in ground water by Kroghet al. 2002.

SedimentsAlcohol ethoxylates were detected in ten sediments of Po riv-er in the range 160–1600 lg/kg (Cavalli et al., 2000).

Alcohol ethoxylates and nonylphenol ethoxylates weredetected in marine sediments of different Spanish coastsand harbours in the range of 37–1300 lg/kg and 8–1080 lg/kg, respectively (Petrovic et al., 2002). More recentlylevels of alcohol ethoxylates within the same order of magni-tude than LAS were found in the Bay of C�diz area (LaraMartin et al., 2006). Alcohol ethoxylates were also detectedin three US river sediments in the range 171–919 lg/kg,most of these were represented by free alcohols (Dyer et al.,2006).

Alkylphenol ethoxylates

1) Laboratory data

Alkylphenol ethoxylates showed poor to moderate minerali-zation rates (20–50% biogas formation) in screening andsimulation tests.

Simulation tests indicated that degradation proceeds viade-ethoxylation to alkylphenol which is poorly degradable.Consequently, high concentrations of nonylphenol (up to

2530 mg/kg DM) were measured in digestors (Tschui andBrunner, 1985).

2) Monitoring data

A complete review of the environmental fate of alkylphenolsand alkyl phenol ethoxylates was published by Guang-GuoYing 2002. APEs are biodegradable in aerobic and anaerobicconditions but aerobic are preferred in fact, concentrationsof alkylphenol ethoxylates in the range <0.1 to 13700 lg/kgwere detected in sediments.

Low concentrations (35–95 mg/kg DM) of nonylphenolin digester sludge have been reported (K�chler, 1995).

Sugar derivatives

1) Laboratory data

Alkylpolyglucosides (C8–-C12-linear alkyl) and glucosidefatty acid esters are well degradable in anaerobic biodegrada-tion screening tests (>60% biogas formation).

Glucose Amides showed high mineralisation rates(>80%) in a 14C-simulation test.

2) Monitoring data

No data available.

Amine oxides

Amine oxides are likely to be anaerobically degradable (1 po-sitive data point – DDAO).

Alkyl ethanolamides

The anaerobic biodegradability of cocomonoethanolamidehas been examined by using the ISO 11734 method. Ulti-mate anaerobic biodegradation reached 81% after 56 daysat 35 8C (Madsen, 2001).

7.3 Cationics – Amphoterics

SurfactantType

Characterisation Test Type Test Conc. Inocu-lumConc.

TestDur-ation


Screen-ing

Simu-lation

mg/lactivematter

mg/lcarbon

g/l days 8C %

Cationics

DTDMAC Dimethyl-di(14C)Stearyl-ammo-nium chloride

(C18)

14C-di-gester

10 *8 *20(as drymatter)

28 35 8.2(CH4) + 6.7(CO2) = 14.9

degradationdue to impurity

?

Steber (1991)

DHTDMA-C

HydrogenatedTallow-based

ECETOC check notknown

35 No degrada-tion

Garcia et al.,(2000)

STAC 14C1-StearylTri-Methyl ammo-nium chloride

(C 18)

14C-di-gester

0.98 -mg/kgsedi-ment

ww-pondsedi-ment500 -ml/l

87 22 0 Federle andSchwab (1992)

CTMAB Cetyltrimethylammonium bro-mide (C 16)

Mod.Shelton &Tiedje

50 50 2–3 gl 60 35 inhibition Battersby andWilson (1989)

MTEAEsterquat

Esterquat ECETOC 50 1 to 5(as drymatter)

42 35 101.1 ± 12.8 Puchta et al.,(1993)



SurfactantType

Characterisation Test Type Test Conc. Inocu-lumConc.

TestDur-ation


Screen-ing

Simu-lation

mg/lactivematter

mg/lcarbon

g/l days 8C %

DEEDMAC Esterquat(C 16–18)

ECETOC *38 1 to 5 60 35 90 Giolando et al.,(1995)

BAHMA Bis (acyloxy-ethyl)-hydroxy-methyl-methylammoniummethosulfate

ISO11374

50 notknown

120 25 24(Low numberpossibly dueto inhibition )

Fraunhofer,(2003)

QFAI Quaternized fattyacid imidazolinemethosulfate

ISO11374

50 notknown

120 35 64 Fraunhofer,(2003)

Ampho-terics

Betaines Coco Betaine ISO11374

100 notknown

84 35 0(Inhibition ?)

(indications ofsubstantial pri-mary degrada-

tion)

Fraunhofer(2003)

CocoAmidoPropylBetaine

ISO11734

50 notknown

56 35 75 Madsen andPetersen,(2001)

CocoAmidoPropylBetaine

ISO11734

100 notknown

84 35 60–80 Also highprimary

degradation(>99%)

Fraunhofer,(2003)

Alkyl am-phoace-tates

Disodium co-coamphodiace-

tate

ISO11734

56 35 78% oftheoreticalgas produc-

tion

Petersen andAndersen,(2004)

Conclusions regarding anaerobic degradability of cationicsurfactants

Cationics

1) Laboratory data

Anaerobic degradability results for cationic surfactants needto be evaluated with caution, since they may be inhibitory toanaerobic metabolism at low levels (i. e. low mg/l range).Nevertheless, some apparent trends can be observed:

y Mono-alkyl or di-alkyl quaternary nitrogen compoundswith straight (C–C) alkyl chains are not anaerobically de-gradable (e.g. TMAC, DTDMAC, etc).

y Esterified mono-alkyl or di-alkyl quaternary nitrogencompounds are biodegradable (e. g. MTEA esterquats,DEEDMAC, etc). After ester hydrolysis, both the fattyacid as well as the alcohol cleavage products can befurther completely mineralized.

y Only a single study is listed on quaternary imidazoliumsalts, which seems to indicate the potential for anaerobicdegradability.

Since only few compounds have been tested, it is diffi-cult to draw general conclusions for cationics with a high de-gree of confidence. The effect on the degradability of factorssuch as the position of the ester bond, the type of substitu-tion on the nitrogen atom, branching, etc., have not beensystematically investigated.

For some categories of cationic surfactants, such a benzal-konium- or ethoxylated quaternary salts no data were found.

These compounds represent, however, only small usage vol-umes as compared to the compounds in the table 7.3.

2) Monitoring data

Some monitoring data for DTDMAC/DODMAC in sludge,sediments and soils are e.g. provided in the EU Risk Assess-ment Report on DODMAC (2002).

There is no indication of DTDMAC being substantiallydegraded in anaerobic sediments, or other anaerobic com-partments.

STAC was not mineralized in anaerobic sediments of alaundromat pond (Federle & Schwab, 1992).

These results are in line with the conclusion from labora-tory tests on these chemicals.

Conclusions regarding anaerobic degradability of amphotericsurfactants

Amphoterics

Laboratory data

The only categories of amphoteric surfactants for which datawere found are betaines and amidobetaines.

y Alkylbetaines: There remains uncertainty about the ulti-mate anaerobic degradability of alkyl betaines. In the re-ported experiment a high degree of primary degradationwas observed, but no apparent mineralization.



y Alkylamidobetaines are anaerobically well biodegradable.y Alkylamphoacetates are also ultimately biodegradable

under anaerobic conditions.

No data were found for other groups of low volume am-photerics such as glycinates, aminopropionates, sultaines, etc.

8 Evaluation of the Relevance of Anaerobic Biodegradation

8.1 Understanding mass fluxes of surfactants in environmentalcompartments

A simplified scheme is used here to illustrate the typicalfluxes of surfactants through a Sewage Treatment Plant(STP), and into different environmental compartments(Fig. 7). The mass fluxes are driven by the following factors:

y The adsorption of surfactants on primary solids, whichwill determine the fractions that go to aerobic and anae-robic zones of a STP, respectively,

y The degradation extent of the surfactant in these twozones taking into account the typical residence time(6 h for the aeration tank and up to 3 weeks for the anae-robic digester),

y The degradation extent in the receiving environmentalcompartment, assuming a typical residence time (1 yearin agricultural soil, 13 years in landfill).

100% sewage treatment, in line with the objectives ofthe EU Urban Wastewater Treatment Directive. (N.B: a di-rect discharge scenario of raw sewage is not included heresince other factors such as ammonium toxicity and oxygendepletion will have a major environmental impact that willoverride any potential surfactant effect). Besides, the scenar-io is still conservative since it does not consider biodegrada-tion on the sewers (Up to 60%; Moreno and Ferrer, 1990and Matthijs and Debaere, 1995).

Statistics from ADEME (1999) and Witte (2001) on thedisposal routes of sewage sludge in the EU were used to de-fine the average fraction of sludge disposed to agriculturalland & forestry (50%), landfill (25%), incineration (20%)and other outlets (5%). The other fraction covers minorroutes such as e. g. composting, pyrolysis and wet oxidation.

To illustrate the general behaviour and fluxes of surfac-tants with different degradability patterns, two approacheswere followed: a modelling approach for three hypotheticalsurfactants and an approach based on monitoring data forreal surfactants. Both approaches are complementary be-cause data availability is extensive for LAS but less so forother surfactant types.

8.1.1 Mass fluxes obtained via modelling for hypothetical surfactants

In the modelling approach Case 1 the surfactant is both aero-bically and anaerobically degradable, whereas in Case 2 thesurfactant is aerobically but not anaerobically degradable.The surfactant in Case 3 is neither aerobically nor anaerobi-cally degradable. It should be noted that a Case 3 surfactantwould not be allowed in the market place today, according tothe EU Detergent Regulation (648/2004/EC), but the exampleis used here to illustrate the effect of biodegradability).

In order to derive values for the various fluxes and toreach conclusions with regard to the amount of surfactantpresent in the various anaerobic compartments the follow-ing working assumptions were made.

y 100% of the surfactant used enter a STP. The potentialdegradation that may have occurred already in the seweris not taken into account.

y 25% of surfactant are removed by adsorption on primarysolids. The remaining 75% go to aerobic treatment,being consistent with a Kd of ca. 2000 l/kg).

y For surfactants that are readily (aerobically) degradable(consistent with a primary degradation rate of 3/h) it isassumed that– 98% are removed in a STP– 95% are removed in landfill– 99% are removed in agricultural soil (consistent with

a primary half life <10 days)y For surfactants that are also anaerobically degradable, it

is assumed that 80% are removed by anaerobic digestiony Since the disposal of sludge to the sea is not allowed any-

more, this route was not included

N.B: As noted in Figure 7, a) Flux 1 does not take intoaccount potential biodegradation in the sewer, and b) someSTPs have the capacity for aerobic stabilisation or compost-



* Flux N06 (sludge for agricultural use) mayundergo aerobic stabilisation prior to use onland

Figure 7 Schematic representation of sur-factant flux through a WWTP and final disposal

ing of the sludge after anaerobic digestion. This processwould remove more than 90% of aerobically degradable sur-factants. As a result the numbers in Table 2 can be seen as aconservative scenario. The purpose of these examples is toillustrate the general behaviour and fluxes of surfactantswith different degradability patterns rather than making arealistic exposure evaluation on a country per country basis.The model demonstrates how similar/different cases 1 and 2are and what the impact is of the anaerobic biodegradationon the fluxes.

From Table 2 it becomes apparent that provided a surfac-tant is readily (aerobically) biodegradable, even if it is notanaerobically degradable (case 2), less than 1% of the surfac-tant released to a STP will reside in the permanently anaero-bic compartments of a landfill, and the occasionally anaero-bic compartment of agricultural lands. For comparison, asurfactant that is both aerobically and anaerobically biode-gradable (case 1) will show equilibrium levels that are5 times lower, approximately.

However, in the case of an aerobically and anaerobicallynon degradable surfactant (case 3), the percentage presentin agricultural land may be estimated to be at least 10% ofthe yearly flux. This is not only 100 times higher than incase 2, but additionally it must be expected that the concen-trations will increase continuously and may eventually causeecotoxicological problems. At the same time, the flux whichmay end up sorbed onto river sediments is more than30 times higher, again with the danger of increasing envi-ronmental concentrations. It is for these reasons that theEU legislation prescribes rapid aerobic degradability, andthat surfactants which are not aerobically degradable arenot in use by the detergent industry.

8.1.2 Mass fluxes based on monitoring data

A similar balance of mass fluxes may be drawn up for con-crete detergent surfactants where monitoring data are avail-able. The environmental fate of LAS has been most exten-sively studied since it is one of the most widely usedsurfactants and easy to determine analytically. Although

monitoring data for other surfactants are available (see ap-pendix), the information does not allow to complete allfluxes.

Giger et al. (1989) investigated surfactant fluxes in theZurich STP and found LAS removal in the primary settlerto be 27%. Berna et al. (1989) doing similar work on aWWTP in Madrid found 16% removal, whereas Prats et al.(1997) working on the WWTP of Alicante found this fractionto be 37%. The discrepancy between the latter case and theother two is explained by the unusually high water hardnessin the Alicante region, and consequently the strongly de-creased solubility of LAS in sewage (cf. 8.2.1). It seems thata removal of 25% of LAS in the primary stage of a STPwould be a fair estimate. This would mean that 75% of theLAS entering a STP will go to aerobic treatment, with thaton primary solids going to the anaerobic digester.

In an anaerobic digester, Prats et al. (1997) found 18%removal of LAS, while Giger et al. (1987) found 20–30%and Osburn (1986) 0–35% removal.

Although degradation tests under strictly anaerobic cir-cumstances indicate that sulfonates are not anaerobically de-gradable, studies on these materials show extensive degrada-tion in case there has been a short exposure to oxygen. Theremovals found in the monitoring studies may therefore re-sult from that fraction of LAS entering the digester whichhad been pre-exposed to aerobic conditions. In addition tothis, in some WWTPs some anaerobic sludge is returned tothe primary settler prior to aerobic composting or dischargeto the chosen sludge disposal route. To simplify matters itseems reasonable to take 15% removal of LAS by anaerobictreatment as a fair estimate. This percentage of removal hasnot been used in the modelling approach.

Prats et al. (1997) found that an aerobic stabilisationstep – where applied – removes a further 65% of the re-maining LAS. The sludge may be disposed of to agriculturalland, to landfill, to the marine environment (practice endedin 1998) or to incineration. The same authors reported LASremoval values of nearly 100% using specific analyticalmethods during composting of anaerobic sludge in a com-mercial composting plant (Prats et al., 1999).

For the LAS fraction going to the aerobic process of asewage treatment plant, Prats et al. (1997) found the amountleaving in effluent to be 0.6% of that entering the STP,equivalent to 99.4% removal by the aerobic treatment. Rapa-port and Eckhoff (1990) found 98% removal by activatedsludge plants although this was slightly lower for tricklingfilter plants. Feijtel et al. (1995) found an average of 99.2%removal when monitoring STPs in the Netherlands. Inagreement with the HERA risk assessment for LAS (HERA,2004) a figure of 99% removal is a reasonable average,hence, resulting in release of 1% to receiving waters andtheir sediments.

Holt et al. (1989) recorded in excess of 98% LAS removalin sludge-amended soils over a year, and calculated a half-life of 7–22 days. Ferrer et al. (1996) found 89.2% removalwithin 62 days giving a half-life of 19.3 days, while K�chlerand Schnaak (1997) determined a 7-day half-life of LAS in afield trial study during the growth season.

Marcomini et al. (1988) monitored a landfill site havingreceived sewage sludge and found at the surface of the sitewhere sludge application was most recent, an LAS concen-tration of 9160 mg/kg while near the bottom this haddropped to 245 mg/kg. This equates to an observed removalof 97% over a 13-year period.

There have been additional monitoring studies on othersurfactant types but not as extensive as those for LAS. Pratset al. (1997) also followed the fluxes of nonionic surfactants



Case 1(%)

Aerobic +Anaer. +

Case 2(%)

Aerobic +Anaer. –

Case 3(%)

Aerobic –Anaer. –

Flux 1 (input) 100 100 100

Flux 2 (to aerator) 75 75 75

Flux 3 (to digester) 25 25 25

Flux 4 (to river) 2.0 2.0 75.0

Flux 5 (from digester) 5 25 25

Flux 6 (to agriculture) 2.5 12.5 12.5

% present in agricultural soilafter 1 year

0.025 0.125 >12.5 *

Flux 7 (to incineration) 1 5 5

Flux 8 (to landfill) 1.25 6.25 6.25

% present in a landfill after13 years

0.0625 0.31 >6.25*

Flux 9 (to other) 0.25 1.25 1.25

* The amounts cannot be accurately estimated because no equilibrium is reached,Hence the given value is considered as a minimum, and a continuous build-up inthis compartment can be assumed.

Table 2 Modelled mass flux of hypothetical surfactants through a STP, andfinal concentrations present in a landfill and in agricultural soil

through a STP. However, whereas in their study of LAS thismaterial was followed specifically by HPLC, nonionics wereanalysed using the Wickbold method (BiAS), a techniquewhich is not very specific and, hence, may also measurenon-ionic materials other than nonionic surfactants.

Field et al. (1995) studied the flux of secondary alkanesulfonates (SAS) through a STP, allowing to construct themass balance in Table 3.

From all these studies approximate conclusions can bedrawn as to the amount of surfactant leaving WWTPs andentering various environmental compartments, but thereare insufficient data available to assess their removal inthese compartments in great detail. It is however possibleto fill these gaps using circumstantial evidence. Since SAS,like LAS, is aerobically but not anaerobically degradable, onecan assume comparable removals in landfill and agriculturalland of 98 and 97%, respectively. Alcohol ethoxylates (AE)are aerobically and anaerobically degradable and so removalsof 99% could be expected in these compartments.

Alcohol ethoxy sulfates (AES), alcohol sulfates (AS), AEand soap were also monitored in Dutch STPs (Feijtel et al.(1995) using a highly specific analytical method. Based onmeasurement of the influent and effluent concentrations,removals of >99% were determined, corresponding to a fluxto rivers of 0.2% (AE), 0.4% (AES), 0.8% (LAS, AS) and0.95% (soap), respectively. It is worth to note that the meas-ured Biochemical Oxygen Demand (BOD) flux to rivers was1.9%, i. e. the surfactants were removed to a greater extentthan the sewage, and at least as efficiently as soap.

From the above monitoring data on these surfactants, amass fluxes table was drawn up (Table 3).

These monitoring data are in general agreement with themodelled scenario of surfactant mass fluxes, indicating agood conceptual understanding of surfactant behaviour inthe environment.

According to the values in Table 3, the level of an anaero-bically degradable surfactant (e.g. AE) in soil after 1 year willbe 6–8 times lower than that of one which is only aerobi-cally degradable (e.g. LAS). It illustrates that surfactantswhich are readily biodegradable under both aerobic andanaerobic conditions should have overall a lower potentialto cause an impact on soils and terrestrial biota.

However, the fact that a surfactant is anaerobically de-gradable does not exclude its presence in anaerobic environ-mental compartments as shown by field monitoring (see8.2.1).

8.2 Impact of surfactants on structure and functionof anaerobic environmental compartments

8.2.1 Speciation and bioavailability of surfactants in anaerobiccompartments

The risk assessment of surfactants typically assumes that so-luble species (usually Na salt; most bioavailable and biode-gradable forms) are present in the environment. In analyti-cal determinations of surfactants in soil, sediment andsludge samples, the soluble form is created during the sam-ple preparation. Nevertheless, in reality surfactants can besubject to complex sorption and speciation processes (RefRISICO).

The speciation of surfactants present in anaerobic com-partments has never been established exactly. The tendencytowards precipitation of surfactants, particularly anionics,when reacting with water hardness (Ca++, Mg++) ions consti-tutes together with adsorption and absorption phenomenathe most important mechanism ruling the presence of sur-factant in environmental solid matrices. Such phenomenamight also occur with cationics as a consequence of the in-teraction with various anionic species.

Understanding speciation of surfactants in anaerobiccompartments poses a scientific challenge, but would be avaluable tool to assess the bioavailability and toxicity behav-iour. There are, for example, several indirect indications ofthe presence of LAS in solid environmental samples underthe form of insoluble Ca–Mg derivatives. While total surfac-tant concentrations are sometimes high, the expected toxi-city is not always observed, probably due to a low bioavail-ability.

The solubility of Ca – LAS homologues in water is verylow. For C12 LAS the solubility at 250 mg/l water hardnessis in the order of 9 mg/l. This noticably influences the acutetoxicity to Daphnia Magna in such a way that, as the precipi-tation progresses, the toxicity of LAS diminishes even athigh nominal concentration (Verge et al., 1997).

A clear relationship between LAS concentrations in pri-mary sludge as well as on anaerobically digested sludgeand water hardness has been observed (Berna, 1989). Acomparison of the mass balance of Ca++ and Mg++ ions inthe raw sewage, treated water and digested sludge indicatesa remarkable accumulation of those ions in the sludge (Ber-na, 1992). While the concentration of Ca++ + Mg++ is nearlyhalf of the amount of Na+ in raw sewage as well as in treatedwater, in the final sludge the concentration of Ca++ + Mg++ isalmost 30 times higher than [Na+].

A remarkable observation is that anaerobic digestion isnot inhibited by the presence of high levels of LAS, evenabove the known inhibitory level of the Na-salt (Painter andMosey, 1992; see also section 8.2.2.1). This is attributed to alimited bioavailability. Hence, speciation aspects are impor-



AE [b] LAS SAS [a]

Flux 2 (to aerator) 44% 75% [b, e, f] 84.4%

Flux 3 (to digester) 41% 25% [b, e, f] 14.6%

Flux 4 (from digester) 14% 21.3% [b, e, g] 16.3%

Flux 5 (to river) 0.2% [d] 1.0% [b, d, h] 0.3%

Flux 6 (to agriculture) 1.6% 7.7% 5.9%

% present in agriculturalsoil after 1 year

0.02% 0.15% [i] 0.12%

Flux 8 (to landfill) 2.1% 9.9% 7.6%

% present in a landfill after13 years

0.02% 0.3% [j] 0.23%

Data in italics and with a shaded background are calculated from assumptionsgiven earlier in the text. The references provide the specific data sources.

[a] Field et al. (1995), [b] Prats et al. (1997),[d] Feijtel et al. (1995), [e] Giger et al. (1989),[f] Berna et al. (1989), [g] Osburn (1986),[h] Rappaport and Eckoff (1990), [i] Holt et al. (1989),[j] Marcomini et al. (1988)

Table 3 Estimated mass flux of high-volume surfactants based on monitor-ing data

Na+ Ca++ Mg++

Raw sewage (mg/l) 63 28 9

Treated water (mg/l) 60 26 8

Sludge (mg/kg) 1,200 21,500 9,000

Table 4 Comparison of concentrations of Na+, Ca++ and Mg++ in rawsewage, treated effluent and in sludge (Berna 1992)

tant to be considered when assessing the realistic ecotoxico-logical impact of surfactants.

The same situation is most likely applicable to all anionicsurfactants. There are several publications showing consid-erably high soap concentrations in anaerobic compartmentssuch as anaerobic sludge (Moreno et al., 1993) and marinesediments (Folke et al., 2003). The anaerobic biodegradationof soap in laboratory scale digesters as well as in full scaleplant is around 80% (Prats et al., 1999). The low availabilityand slow dissolution of Ca–Mg soap account for the lessthan complete degradation.

While LAS present in those environmental compart-ments can only be linked to its use in man-made detergentproducts, the origin of soap, at least in compartments show-ing extremely high concentrations like marine sediments, isnot yet clearly understood. The possibility of biochemical re-actions involving fats and oils from a variety of (natural) ori-gins taking place in the environment cannot be excluded.For risk assessment purposes the total bioavailable concen-tration of soap should be considered regardless of its origin.

8.2.2 Impact on wastewater treatment and aquatic environments

8.2.2.1 LAS and other anionic surfactants in anaerobic digesters

There is relatively little published information available onthe inhibitory effects of surfactants on anaerobic digestion.Bruce et al. (1966) reported results obtained from laboratorydigesters fed on a daily basis with surfactant-free raw sludgewhich had been amended with different concentrations ofsurfactant. LAS (C 10–C 13, sodium salt) at a concentrationof 15 g/kg dry solids was not inhibitory; although biogas pro-duction was reduced over the first 24 h after feeding, totalgas evolution after 48 h was similar to that in controls fedwith surfactant-free sludge only. At LAS levels between 15and 20 g/kg dry solids the reliability of digestion was im-paired and more serious inhibition occurred at higher con-centrations. These findings were confirmed by Metzner(1998), who found inhibitory effects on gas production atca. 1.1 g LAS/l resp. 22.5 g LAS/kg dry solid residue at adry solid concentration of 50 g/l. Similar results were ob-tained for SAS by Bruce et al. (1966). C 16 – C 18 alcohol sul-fate from natural raw materials as well as a petrochemicalbased primary alcohol sulfate had both only a minor short-term inhibitory effect on gas production at concentrationsup to 40 g/kg dry solids. At the end of the 18 day incubationperiod, gas production in the test digesters was higher thanin the controls and 90% of the surfactants had been de-graded (MBAS removal). Nonylphenol ethoxylate had onlya slight inhibitory effect on gas production at levels up toeven 200 g/kg dry solids.

Published concentrations of surfactants in digestersludge show that the majority of the data are for LAS, withaverage levels around 5 g/kg dry solids (Schowanek et al., inpress). Hence, LAS levels in sewage sludge at most sewagetreatment works are below inhibitory levels. The toxicity ofLAS to anaerobic methanogenic bacteria was studied usingstandard methods (Sanz et al., 1999). The experiments wereconducted on UASB (Upflow Anaerobic Sludge Blanket)type reactors with a working volume of 315 mL. The 50%inhibition concentration (IC50) of the methanogenic activitywas reached at concentrations higher than 25 mg/l of LAS(in the range of 40 to 150 mg/l) in all cases. These figuresare in agreement with other literature data, and are higherthan what is typically found in domestic wastewater.

The anaerobic digestion of AS-rich textile wastewater(e.g, from cotton desizing) was tested in batch experiments

(Feitkenhauer & Meyer, 2002). In the first part of the studyC10 AS was used as model to study the influence of concen-tration in the potential inhibitory effects. The starch hydro-lysis was inhibited at concentrations above 65 g AS/kg CDW(cell dry weigh). In the second part of the study the inhibi-tory effects was shown to be higher for medium and longchain AS. The concentrations above which the inhibitory ef-fects are observed are high. In domestic STPs no inhibitionattributed to AS has been reported.

Similar effects have been found before in experimentswith long chain fatty acids (soaps; Angelidaki and Ahring,1992). More recently, the performance of laboratory scaleUASB and DASB reactors in the anaerobic biodegradationof synthetic wastewater containing sodium oleate and so-dium stereate was tested. The major problem encounteredwas the adsorption of soap onto the solid support, impedingtherefore biomass-substrate contact and affecting the overallbiodegradation performance (Miranda et al., 2006).

8.2.2.2 Other surfactants in anaerobic digesters

The concentration of linear alcohol ethoxylate (AE) in diges-ter sludge is around 0.5 g/kg dry solids, i. e. an order of mag-nitude lower than LAS (Klotz, 1998). AE concentrations of acomparable level were reported by Cavalli and Valtorta(1999) and by Bruno et al. (2002) (<700 and 312 mg/kg, re-spectively) while Matthijs et al. (2004) detected AE concen-trations in the range of <22–468 (mean 168) mg/kg in thedigestor sludges from several European STPs. These AEconcentrations measured in practice can be compared withthe AE concentrations shown to be non-inhibitory in anaero-bic degradation screening tests in the laboratory (cf. data inchapter 7). The unrealistically high surfactant to biomass ra-tio in these tests represents a worst case. Therefore inhibi-tory effects in practice can be excluded if the surfactant con-centration in practice does not significantly exceed the (non-inhibitory) screening test concentration. A positive screen-ing test result was obtained by Fraunhofer (2003) with a testconcentration of 120 g AE/kg sludge DM, i. e. more than2 orders of magnitude higher than realistic AE concentra-tions. Hence, inhibitory effects of AE on the function ofanaerobic digestion can be excluded.

Bruno et al. (2002) reported concentrations of other sur-factants in digester sludge of the Roma Nord STP (Italy), re-ceiving mainly dometic sewage. The following concentra-tions were determined in 24-h composite samples ofdigested sludge: 50 mg AS /kg dw and 69 mg AES /kg dw.These can be compared with non-inhibitory test concentra-tions of 96 g/kg for AS (Salanitro & Diaz, 1995) and ca.17 g/kg for AES (Fraunhofer, 2003), respectively, underlin-ing again that the realistic concentration of these surfactantsin digester sludges are well below the inhibition thresholdcritical for the functioning of anaerobic digesters.

Despite the widespread use of anaerobic digestion, theoperation of these installations is not known to be disturbedby the presence of surfactants.

8.2.2.3 Septic tanks and decentralized treatment systems

Septic tanks are onsite (pre-) treatment systems for blackwater or unsettled domestic sewage. They are not consideredanymore state-of-the-art systems for sewage treatment butnevertheless are still in use in Europe and the US, particu-larly in rural areas. Septic tank effluents can be sent to asewer system, a ditch, or may be allowed (not generally) topercolate into the soil (i. e. into a tile field). The septic tankitself is an anaerobic/microaerobic system while the tile



field in the immediate vicinity of the tank, is usually astrictly anaerobic environment. Surfactants released via aseptic tank system are therefore exposed first to an anaero-bic/microaerobic compartment without dilution. BOD/CODremoval is around 50% on average (LEAF, 2006).

The main function of a septic tank is to act as a settlingtank for the solids in domestic sewage, and needs to be per-iodically emptied (e.g. once per year). This sludge is treatedelsewhere, usually in a nearby treatment plant, or may be(not generally allowed) disposed to land. The water phasein the tank has a residence time up to 5 days. The sludge atthe bottom of the tank undergoes an anaerobic decomposi-tion, with the formation of biogas and H2S. This digestion isthe second function of a septic tank. Since the temperaturein the tank is close to that in the surrounding soil (<20 8C),digestion typically occurs at a slow rate.

Most surfactants and detergent formulations have beentested for septic tank safety, i. e. when discharged under nor-mal usage condition they will not disturb the functioning ofthe septic tank. The protocol for such tests includes a para-meter related to anaerobic digestion and settling rate underrealistic circumstances. This is mostly done in the labora-tory, but some validation on real septic tanks has been per-formed. Also surfactants which are not anaerobically degrad-able can be safely used in septic tanks. Data to support thisare available within the Industry member companies. Somestudies are also available in the literature on the fate and be-haviour of non-degraded surfactants leaving the tank into atile field (McAvoy et al., 1994; Shimp et al., 1994; McAvoyet al., 2002).

A recent trend in Europe is the increase of so-called de-centralized sewage treatment plants, i. e. small installationsdesigned for individual houses or small communities (<100inhabitant equivalents). They offer a better treatment perfor-mance than septic tanks. In the ERASM-sponsored LeAF Re-port (2006) an overview is given of the incidence of technol-ogy for small and decentralized sewage treatment plants onthe EU market. In contrast to septic tanks, however, most ofthese systems are based on aerobic processes. As a conse-quence, they will not be discussed further in this review.

8.2.2.4 Anaerobic river and lake sediments

Freshwater sediments are usually anaerobic as of as few mmor cm depth, and stratified in terms of redox potential andthe metabolic processes that occur within these zones. Inrivers, the structure of the sediment will be affected by thevelocity and turbulence of the river flow, which in turncauses the transport of sediment particles either by suspen-sion or by sliding and rolling (Press and Siever, 1986). Anae-robic freshwater sediments play an important role in the de-composition of organic matter, and in the cycling ofnitrogen, sulphur and phosphorus. These processes will bedependent on the redox potential of the sediment, and in-clude nitrate reduction to either ammonia or nitrogen, sul-fate reduction to sulphide and methanogenesis (Jones,1982). Sediments can also act as sinks for chemicals whichsorb to sediment particles or organic matter. The sedimentis also home to a wide range of fauna, flora and micro-or-ganisms, and perturbation of these organisms can have adetrimental effect of the structure and function of the eco-system (e.g. Burton and MacPherson, 1995).

Available measured LAS data in freshwater sedimentswere reviewed by Cavalli et al., (2000). Typical LAS valuesin sediments below sewage outfalls were found in the 0.5–5.3 mg/kg range with an arithmetic mean of 2.9 mg/kg (12records, SD = €1.9). Homologue distributions were also

measured for some river sediment samples (Cavalli et al.,2000) and the corresponding fingerprint was found to besimilar to that of sludge and soils.

The occurrence of alkyl sulfates and alkyl ethoxysulfatesin US river sediments has been reported by Sanderson et al.(2006). Combined levels of AS/AES were detected in a pre-sumed lower exposed sites in the range of 0.025–0.034 mg/kg (dw) while a presumed higher exposed sitehad combined levels of 0.117 mg/kg dw.

Little information of the toxicity of LAS (or other surfac-tants) to anaerobic, sediment bacteria could be found. Jen-sen et al. (2001) report an EC10 (= surrogate of a NOEC) of5 mg LAS/kg soil for the anoxic process of microbial iron re-duction. Data on the toxicity of LAS to aerobic bacteria(EC50 = 183 g/kg dw of activated sludge) (HERA, 2004) andanaerobic digestion (EC0 > 15 g/kg, see 8.2.2.1) indicate thatthe levels of LAS in river sediments are not expected to inhi-bit microbial processes.

As regards higher organisms in sediments, Pittinger etal. (1991) reported a NOEC = 319 mg/kg dry solids for thelarvae of a benthic organism, Chironomus riparius (midge).In addition, subchronic toxicity tests with mussels (Unioelongatulus), and worms (Lumbriculus variegatus, Caenor-habditis elegans) revealed NOEC values of >200, 81 and100 mg/kg, repectively (HERA, 2004). Also these data showclearly that the concentration of LAS in freshwater sedi-ments is at least one order of magnitude below the levelswhich might impact freshwater organisms.

Considering the prominent role of LAS in terms of theusage volume in detergents and an ecotoxicological profilesimilar to other relevant surfactants used in detergents, itcan be assumed that the above conclusion on LAS is also val-id for the other detergent surfactants.

8.2.2.5 Agricultural soil (pasture and cropland)

About 50% of the European sewage sludges are disposed onagricultural land (Schowanek et al., 2004). A framework forperforming an ecological risk assessment for chemicals inagricultural soil is provided in e.g. the EU TGD (2003) andby ILSI-Europe (Schowanek et al., 2004). According to theEuropean Technical Guidance documents (EU-TGD, 2003)for sludge application to agricultural soil, an application rateof 0.5 kg sewage sludge DM/m2 per year is assumed(= 5000 kg/ha.year). This sludge is mixed into 20 cm of top-soil, with an average density of 1.3 tonne/m3. The EU TDGapproach also considers a 30-day time-weighted averagingwhile calculating PECsoil, in order to account for the fact thatseeding/planting typically occur a few days/weeks after ap-plication of the sludge.

Assuming an avarage European LAS concentration inanaerobic sludge of 5000 mg/kg, (Schowanek et al., 2007),the predicted initial soil LAS concentration is around10 mg/kg immediately after sludge application3. Aerobicconditions prevail in sludge-amended soils, as shown by sev-eral studies which demonstrated that LAS disappears withhalf lives of 7–30 days (mean value *10 days) (De Wolf,1998; Cavalli, 1998; Schowanek et al., in press). Due to theaveraging time concept the LAS concentration in soil (PEC-soil) as used in EU risk assessment is less than half theexpected initial value of 10 mg/kg, i. e. below 5 mg/kg.

No field monitoring has been performed to determinePEC soil for LAS or any other surfactants under the condi-



3 This value is often lower since (anaerobic) sewage sludge is often stored andstabilized for a few weeks or months before use in agriculture. During thisperiod aerobic biodegradation will lower the surfactant concentration.

tions (dosage, mixing, time average, . . . etc.) described in theEU TGD. Nevertheless, many monitoring studies on LASconfirm the general degradation behaviour of the molecule,and importantly, the absence of accumulation behaviourover time.

ILSI-Europe has done an extensive review and probabil-itic risk assessment for LAS in sewage sludge used on agri-cultural soil, looking at a variety of relevant endpoints (Scho-wanek et al., 2004, Schowanek et al., in press). This studyconcludes that the presence of LAS in sludge poses no riskto agricultural soil and the foodchain, up to very high levels(i. e. 55 and 27.5 g LAS/kg sludge for a homogeneous mix-ing and injection scenario, respectively). Such LAS levelsare not reached in European sewage sludge.

Laboratory studies carried out on sludge stored on com-pacted soil floors demonstrate that LAS penetrated, evenafter an extended time period of 1 year, only to a minimumdegree into the soil. Most of the LAS migration was found inthe 3 cm thick soil layer. No LAS was present in the waterpercolating through the soil. Leaching of LAS from thesludge amended soil to ground water should not be expectedto happen (Figge, 1991).

The leaching potential of LAS in agricultural soils hasalso been investigated in soils irrigated with water contain-ing measured LAS concentrations. No LAS was detected be-low the top 10 cm soil layer after one year of periodical irri-gations.

For surfactants other than LAS (i. e. AS, AES, SAS, soap,esterquats and AE (in preparation), the respective HERA re-ports (www.heraproject.org) provide a risk assessment basedon the available terrestrial toxicicity data. This provides aconsistent picture that at current usage volumes none ofthese surfactants gives rise to concern for the terrestrial en-vironment, regardless whether they are anaerobicically bio-degradable or not.

It is sometimes suggested that surfactants, applied withthe use of sewage sludge in agriculture, might increase themobility of organic micropollutants. These effects, however,were only detected at very high surfactant concentrationswhich are not relevant in the agricultural practice (K�chler,1995). This can also be explained by the fact that surfactantsapplied via sewage sludge occur in soil in a form adsorbed tosludge particles.

8.2.2.6 Landfills for sludge

Sanitary landfilling is widely used to dispose of diverse typesof waste, including digested and dewatered sludge. How-ever, over the last years the trend in Europe has been to-wards decreased landfilling and a corresponding increasein incineration (Schowanek et al., 2004). Well run landfillsare isolated from the surrounding environment (ground-water, air, rain). This isolation is accomplished with a bot-tom liner system (clay, plastic foils) and daily covering ofsoil.

Conditions inside a landfill may vary spatially and tem-porarily between aerobic, anoxic and strictly anaerobic con-ditions. Landfills are not designed to break down waste,merely to bury it. However, for organic wastes with asuffi-cient moisture content like sludges from wastewater treat-ment plants a landfill may behave similarl to a bioreactorproviding suitable conditions for anaerobic bacteria. Thiswill lead to the conversion of organic wastes into organicacids and ultimately into methane and carbon dioxide (bio-gass), which is often recuperated for energy generation.

Since landfills are not normally opened there is onlyscarce information about surfactant concentrations in land-

fills. One study (Marcomini, 1989) gives data on TPS (Tetra-propylenebenzene Sulfonate) and LAS for four Germanlandfills, which were used as deposits for digested sludgefor up to 30 years. The switch fromTPS to LAS can be usedas a time marker to help to set the age of the landfill. Con-sidering that LAS consumption, and consequently its inputto landfills, has only slightly increased along the years takeninto account, it appears that, with a residence time of13 years, the LAS concentration decreased from about 9 g/kg to 0.2 g/kg in one landfill. This decrease may rather beindicative of anoxic conditions (oxygen limitation) in the re-spective landfill allowing a slow aerobic degradation of LAS.

There is no evidence of disturbance of landfill function-ing by surfactants. Lack of anaerobic biodegradability is notconsidered relevant in landfills.

8.2.2.7 Marine sediments

The structure and function of marine sediments are compar-able similar to that of freshwater sediments, although thereare some key differences. In estuarine and coastal areas, se-diments are usually anaerobic below the surface few mmbut pelagic sediments can be oxic at sediment depths >1 m(Jørgensen, 1982). The stratification of anaerobic decompo-sition and nutrient cycling processes with decreasing sedi-ment redox potential are similar to freshwater sediments.However, a big difference is the abundance of sulfate in sea-water which means that sulfate reduction replaces methano-genesis as the terminal stage in the anaerobic oxidation oforganic matter. In coastal waters this process can convert asmuch organic matter to CO2 as do aerobic organisms inaerobic environments.

Monitoring data for concentrations of surfactants in es-tuarine or marine sediments are scarce, and those data avail-able are mostly confined to LAS. In their review on the en-vironmental safety of LAS, Painter and Zabel (1988)reported levels of 5–17 mg/kg dry solids in estuarine sedi-ments close to the outfalls of probably untreated sewage.LAS was undetectable further out to sea. The same trendwas observed in the study carried out in sediment samplesfrom the Baltic proper and Little Belt (Denmark) by Folkeet al. High concentration of LAS was measured in the innerpart of the shipping port (BABS, the branched alkylbenzeneused in the 50’s and early 60’s, was even detected). As sedi-ments became cleaner with increasing distance from theshipping ports, the levels of LAS decreased to near or belowthe detection limit (0.1 mg/kg sediment DM) indicatingtherefore that Danish marine sediments are not generallycontaminated with LAS. The aforementioned study also in-cluded the monitoring of soap concentrations. The concen-tration found in relatively uncontaminated sediment was15000–40000 times higher than that of LAS. High concen-trations of soap in sediment (1000–2000 mg/kg DM) wereobserved in areas where LAS could not be detected(<0.05 mg/kg DM). The origin of all detected soap couldnot be pinpointed, since soaps can also be formed by naturalprocesses.

The variation with depth of the concentrations of LASand of the long-chain sulfophenylcarboxilic acids (SPCs – re-sulting from LAS biodegradation) was determined in coastalsediments (Le�n et al., 2001). The average surface layer con-centration (0–8 cm depth) of total LAS varied a lot depend-ing on the sample point, More than 100 ppm were detectedin the vicinity of the discharge point on non treated waste-water (this is considered as non representative data) whileLAS levels ranged from values of several 10’s of ppb to fewppm were found in cleaner areas. The vertical profile of LAS



concentration showed a sharp reduction with depth whileSPCs levels increased. The presence of SPCs in sedimentlayers considered to be anoxic, together with the disappar-eance of LAS, led the authors to postulate the existence ofsome kind of mechanism for LAS degradation under anoxicconditions in the marine environment.

The same research group from the University of C�dizstudied further the distribution of LAS as well as AES inthe sediments taken from the same coastal area of Spain(Gonz�lez Mazo et al., 2005). A decrease in LAS concentra-tions was found compared to past studies in one of the sam-pling areas due to the entry into service of a STP six monthsbefore the time of the sampling. With respect to the AESconcentration levels values ranging between 150 and400 ppb were reported. These concentrations are lower thanthose for LAS, especially near the effluent outlet, mainly dueto the lower AES consumption in Spain. More recently, anadditional monitoring study at the Bay of C�diz was carriedout, this time extending the analysis with AEs (Lara Mart�net al., 2005). Concentrations were up to 637 ppb for LAS,861 ppb for AEs and 125 ppb for AES in surface sediments.It is worth noting that concentration of AE is slightly higherthan that of LAS, despite the fact the AE is known to be bio-degradable under both aerobic and anaerobic conditions.The main trend along the sediment columns is towards aconcentration decrease for of all these compounds withdepth.

There is a general paucity of data regarding the toxicityof surfactants to marine sediment dwelling organisms. Thisis an area that ERASM technical committee has started toaddress and information should be available in the comingyears.

References

Ahel, M. and Giger, W.: Determination of alkylphenols and alkylphenol mono-anddiethoxylates in environmental samples by high performance liquid chromato-graphy. Anal. Chem. 57 (1985) 1577–1583.

Ahel, M., Giger, W. and Koch, M.: Behaviour of alkylphenol polyethoxylate surfac-tants in the aquatic environment – I. Occurrence and transformation in sewagetreatement. Wat. Res 28 (1994) 1131–1142.

Ahel, M., Schaffner, C. and Giger, W.: Behavior of Alkylphenol ethoxylate surfactantsin the Aquatic environment – III. Occurrence and elimination of their persistentmetabolites during infiltration of river water to groundwater. Wat. Res.: 30(1996) 37–46.

AISE/CESIO monitoring on European sludges.Amano, K., Fukushima, T. and Nakasugi, O.: Diffusive exchange of linear alkylben-

zenesulfonates (LAS) between overlying water and bottom sediment. Hydro-biologia 235/236 (1992) 491–499.

Angelidaki, I. and Ahring, B. K.: Effect of long-chain fatty acids on thermophilicanaerobic digestion. Appl. Microbiol. Biotechnol. 37 (1992) 802–812.

Angelidaki, J., Hagensen, F. and Ahring, B. K.: Anaerobic transformation of LAS incontinuous stirred tank reactors treating sewage sludge. In Proceedings of the5th World Surfactant Congress. Comite Européen des Agents de Surface etLeurs Intermediares Organiques. Milano, Italy 2000.

Angelidaki, I., Torang, L., Waul, C. M. and Schmidt, J. E.: Anaerobic bioprocessing ofsewage sludge, focusing on degradation of linear alkylbenzene sulfonates(LAS). Water Science and Technology 49 (10, Wastewater Sludge as a Re-source), (2004) 115–122.

ASTM E 1196-87. Standard Method for Determining the Anaerobic BiodegradationPotential of Organic Chemicals. (1987) 738–742.

Austermann-Haun, U., Kroiss, H., Fischer, P., Joerdening, H.-J., Meyer, H., Pascik, I.,Rosenwinkel, K.-H., Seyfried, C. F., Strotmann, U. J., Svardal, K., Temper, U.,Weiland, P., Winter, J. and Metzner, G.: Fünften Arbeitsbericht der ATV-DVWK-Arbeitsgruppe IG 5.1. KA – Abwasser, Abfall, 51 (2004) 997–1001.

Battersby, N. S. and Wilson, V.: Evaluation of a serum bottle technique for assessingthe anaerobic biodegradability of organic chemicals under methanogenic con-ditions. Chemosphere 17 (1988) 2441–2460.

Battersby, N. and Wilson, V.: Survey of the anaerobic biodegradation potential oforganic chemicals in digesting sludge. Appl. Environ. Microbiol. 55 (1989)433–439.

Baumann, H. and Schefer, W.: Simple test system to judge the anaerobic biode-gradability of organic substances. Textilveredlung 25 (1990) 248.

Baumann, U. and Mueller, M. T.: Determination of anaerobic biodegradability with asimple continuous fixed bed reactor. Wat. Res. 31 (1997) 1513–1517.

Berna, J. L., Ferrer, J., Moreno, A., Prats, D. and Ruiz, F.: The fate of LAS in the en-vironment. Tenside Surf. Det. 26 (2) (1989) 101–107.

Berna, J. L., Moreno, A. and Ferrer, J.: An assessment of the ultimate biodegradationof LAS. Proceedings of CESIO World Surfactants Congress, London (UK), 3(1992) 59–72.

Bester, K., Theobald, N. and Schroeder, H. Fr.: Nonylphenols, nonylphenol-ethoxy-lates, linear alkylbenzenesulfonates (LAS) and bis (4-schlorophenyl)–sulfonein the German Bight of the North Sea. Chemosphere. 45 (2001) 817–826.

Birch, R. R., Biver, C., Campagna, R., Gledhill, W. E., Pagga, U., Steber, H., Reust, H.and Bontinck, W. J.: Screening chemicals for anaerobic biodegradability. Che-mosphere 19 (1989) 1527–1550.

Birch, R. R., Geldhill, W. E., Larson R. J. and Nielsen, A. M.: Proc. 3rd CESIO Interna-tional Surfactants Congress and Exhibition, London, (1992) 26–33.

Bouwer, E. J. and McCarty, P. L.: Transformation of halogenated organic compoundsunder denitrification conditions. Appl. Environ. Microbiol. 45 (1983) 1295–1299.

Bruce, A. M., Swanwick, J. D. and Ownsworth, R. A.: Synthetic detergents andsludge digestion: Some recent observations. The Journal and Proceedings ofthe Institute of Sewage Purification 5 (1966) 427–447.

Brunner, P. H., Capri, S., Marcomini, A. and Giger, W.: Occurrence and behaviour oflinear alkylbenzene sulfonates, nonylphenol, nonylphenol mono-and nonyl-phenol diethoxylates in sewage and sewage sludge treatment. Wat. Res. 22(1988) 1465–147.

Bruno, F., Curini, R., Di Corcia, A., Fochi, I., Nazzari, M. and Saperi, R.: Determinationof surfactants and some of their metabolites in untreated and anaerobicallydigested sewage sludge by subcritical water extraction followed by liquidchromatography-mass spectrometry. Environm. Science & Technology 36(2002) 4156–4161.

Burton Jr., G. A. and MacPherson, C.: Sediment toxicity testing issues and methods.In Handbook of Ecotoxicology, eds. D. J. Hoffman, B. A. Rattner, G. A. Burton Jr.and J. Cairns Jr., 70–103. CRC Press Inc. (1995).

Cantero M., Soledad, R. and Pérez-Bendito, D.: Determination of non-ionic poly-ethoxylated surfactants in sewage sludge by coacervative extraction and iontrap Liquid chromatography-mass spectrometry. J. Chromatogr. A. 1046(2004) 147–153.

Carlsen, L., Metzon, M.-B. and Kjelsmark, J.: Linear alkylenzenene sulfonatee (LAS)in the terrestrial environment. Sci. Tot. Environ. 290 (2002) 225–230.

Casellato, S., Aeillo, R., Negrisolo, A. and Seno, M.: Long-term experiment on Bran-chiura sowebyi Beddard (oligochaeta, tubificidae) using sediment treated withLAS (Linear Alkylbenzene Sulfonate). Hydrobilogia 232 (1992) 169–173.

Cavalli, L., Gellera, A. and Landone: LAS removal and biodegradation in a wastewater treatment plant. Env. Tox. Chem. 12 (1993) 1777–1788.

Cavalli, L. and Valtorta, L.: Surfactants in sludge-amended soil, 1998, XXVIII CED,Barcelona, May 6–8 (1998).

Cavalli, L., Cassani, G., Viganò, L., Pravettoni, S., Nucci, G., Lazzarin, M. and Zatta,A.: (2000). Surfactants in sediments. Tenside Surf. Det. 37 (5) (2000) 282–288.

Cohen, A., Klint, K., Bowadt, S., Persson, P. and Jonsson, J. A.: Routine analysis ofalcohol and nonylphenol polyethoxylates in wastewater and sludge using liquidchromastography-electrospray mass spectrometry. J. of. Chrom. A, 927 (2001)103–110.

Compagnion, D. and Nyns, E.-J.: Biomethanation in developing countries. In Bio-technology Vol. 8. Microbial Degradations, ed. W. Schönborn VCH, (1986)239–267.

Chalaux, N., Bayona, J. M. and Albaiges, J.: Determination of nonylphenols as pen-tafluorobenzyl derivatives by capillary gas chromatography with electron cap-ture and mass spectromentric detection in environmental matrices. J. Chro-matogr. 686 (1994) 275–281.

Chien Chih-Ching, Leadbetter, E. R. and Godchaux III, W.: Sulfonate-sulfur can beassimilated for fermentative growth, FEMS Microbiology Letters, 129 (1995)189–194.

DelValls, T. A., Forja, J. M. and Gomez-Parra, A.: Seasonality of contamination, toxi-city, and quality values in sediments from littoral ecosystems in the Gulf ofCadiz (SW Spain). Chemosphere. 46 (2002) 1033–1043.

DiCorcia, A., Samperi, R., Bellioni, A., Marcomini, A., Zanette, M., Lenr, K. and Ca-valli, L.: LAS Pilot study at the “Roma-Nord” sewage treatment plant and in theTiber river. Riv. Ital. Sost. Grasse 71 (1994) 467–475.

Denger, K., Kertez, M. A., Vock, E. H., Schon, R. and Cook, A. M.: Anaerobic desul-fonation of 4-tolylsulfonate and 2-(4-sulfophenyl) butyrate by Clostridium spp,Appl. Environ. Microbiology 62 (1996) 1526–1530.

Denger, K. and Cook, A. M.: Assimilation of sulfur from alkyl- and arylsulfonates byClostridium spp, Arch. Microbiol. 167 (1997) 177–181.

Denger, K., Lane, H. and Cook, A. M.: Anaerobic taurine oxidation: a novel reactionby nitrate-reducing Alcaligenes spp, Microbiology 143 (1997) 1919.

Denger, K. and Cook, A.: LAS bioavailable to anaerobic bacteria as a source of sulfur,Appl. Environ. Microbiol. 86 (1999) 166–168.

Deuser, W. G.: Organic carbon budget of the Black Sea. Deep Sea Research 18(1971) 995–1004.

De Wolf, W. and Feijtel, T.: TRA for LAS in sludge-amended soil, Chemosphere, 36(1998) 1319–1343.

DiCorcia, A., Samperi, R., Bellioni, A., Marcomini, A., Zanette, M., Lenr, K. and Ca-valli, L.: LAS Pilot study at the “Roma-Nord” sewage treatment plant and in theTiber river. Riv. Ital. Sost. Grasse 71 (1994) 467–475.

Dyer, S. D., Sanderson, H., Waite, S. W,. Compernolle, R. V., Price, B., Nielsen, A. M.,Evans, A., Decarvallo, A. J., Hooton, D. J. and Sherren, A. J.: Assessment of al-cohol ethoxylate surfactants and fatty alcohols mixtures in river sediments andprospective risk assessment. Environ. Mon. & Assess. 120 (2006) 45–63.

Eadsforth, C. V., Sherren, A. J., Selby, M. A., Toy, R., Eckhoff, W. S., McAvoy, D. C. andMatthijs, E.: Monitoring of environmental fingerprints of alcohol ethoxylates inEurope and Canada. Ecotoxycol. and Environ. Saf. 64 (2006) 14–29.

ECETOC. (1988). Evaluation of anaerobic biodegradation. Technical Report 28.EEA (European Environment Agency). Urban wastewater treatment. ETC/WTR

based on Member States data reported to OECD/EUROSTAT Joint Question-naire 2002.

Eichhorn, P., Lopez, O. and Barcelo, D.: Application of liquid chromatography-elec-trospray-tandem mass spectrometry for the identification and characterisationof linear alkylbenzene sulfonates and sulfophenyl carboxylates in sludge-amended soils. Journal of Chromatography A. 1067(1–2) (2005) 171–179.



EPA EPA 40 CFR Ch. 1 (7-1-88 Edition) §796.3140: Anaerobic biodegradability oforganic chemicals. (1988) 568–572.

EU-TGD. Technical Guidance Documents (TGD) in support of EU Directive 93/67EEC and EU Regulation 1488/94 published by the EU Chemical Bureau, 1996.

Federle, T. W. and Schwab, B. S.: Mineralization of surfactants in anaerobic sedi-ments of a laundromat wastewater pond. Wat. Res. 26 (1992) 123–127.

Fernandez, P., Alder, A. C., Suter, M. J. F. and Giger, W.: Determination of the qua-ternary ammonium surfactant ditallowdimethylammonium in digested sludgesand marine sediments by supercritical fluid extraction and liquid chromatogra-phy with post-colums ion-pair formation. Anal. Chem. 68 (1996) 921–929.

Feitkenhauer, H. and Meyer, U.: Anaerobic Digestion of alcohol sulfate (anionicsurfactant) rich wastewater –batch experiments. Part I: Influence of the sur-factant concentration. Bioresource Technology 82 (2002) 115–121.

Feitkenhauer H. and Meyer U.: Anaerobic Digestion of alcohol sulfate (anionic sur-factant) rich wastewater – batch experiments. Part II: Influence of the hydro-phobic chain length. Bioresource Technology 82 (2002) 123–129.

Feijtel, T. C. and van den Plassche, E.: Environmental Risk Characterisation of fourMajor Surfactants used in the Netherlands, Report No 679101025 by NVZ –RIVM (1995).

Ferrer, J. de, Moreno, A., Vaquero, Ma. T. and Comellas, L.: Monitoring of LAS (Lin-ear Alkylbenzene Sulfonate) in direct discharge situations: untreated sewageand on sludges amended soils, CESIO 4th World Surfactant Congress, Barcelo-na (Spain) (1996).

Field, J., Miller, D. J., Field, T. M., Hawthorne, S. B. and Giger, W.: Quantitative de-termination of sulfonated aliphatic and aromatic surfactants in sewage sludgeby ion-pair/supercritical fluid extraction and derivatization GC/MS, Anal. Chem.64 (1992) 3161–3167.

Field, J. A., Field, T. M., Poiger, T., Siegrist, H. and Giger, W.: Fate of secondary alkanesulfonate surfactants during municipal wastewater treatment. Wat. Res. 29(1995) 1301–1307.

Figge, K.: NATEC report, Fate of LAS in digested sewage sludge deposits, February 1991.Folke, J., Cassani, G., Ferrer, J., López, I., Karlsoon, M. O. and Willumsen, B.: Linear

Alkylbenzene Sulfonates, Branched Dodecylbenzene Sulfonates and SoapAnalysed in Marine Sediments from the Baltic Proper and Little Belt. TensideSurf. Det. 40 (1) (2003) 17–24.

Fraunhofer. Anaerobic biodegradation of detergent surfactants. Final report to theEuropean Commission Enterprise Directorate-General, (July 2003). FraunhoferInstitut Umwelt-, Sicherheits-Energietechnik UMSICHT, Oberhausen/Germany2003.

Garcia, M. T., Campos, E., Illan, P., Dalmau, M. and Sanchez-Leal, J.: Study of theanaerobic biodegradation of homologs and pure isomers of linear alkylben-zene sulfonates in discontinuous digesters. Comunicaciones presentadas a lasJornadas del Comite Espanol de la Detergencia 34 (2004) 147–160. ComiteEspanol de la Detergencia, Tensioactivos y Afines.

Garcia, M. T., Campos, E., Ribosa, I., Latorre, A. and Sanchez-Leal, J.: Anaerobic di-gestion of linear alkyl benzene sulfonates: biodegradation kinetics and meta-bolite analysis. Chemosphere 60 (11) (2005) 1636–1643.

Garcia, M. T., Campos, E., Sanchez-Leal, J. and Ribosa, I.: Effect of linear alkylben-zene sulfonates (LAS) on the anaerobic digestion of sewage sludge. WaterResearch 40 (2006) 2958–2964.

Garcia, M., Teresa, C., Encarna, D., Manel, I. P. and Sanchez-Leal, J.: Inhibition ofBiogas Production by Alkyl Benzene Sulfonates (LAS) in a Screening Test forAnaerobic Biodegradability. Biodegradation, 17 (1) (2006) 39–46.

Giger, W., Brunner, P. H. and Schaffner, C.: 4-Nonylphenol in Sewage sludge: Accu-mulation of Toxic Metabolites from Non-ionic Surfactants. Science 225, vol. 10,August. (1984) 623–625.

Giger, W., Brunner, P. H., Ahel, M., McEvoy, J., Marcomini, A. and Schaffner, C.: Or-ganic detergent components and their degradation products in wastewaterand sludge. Gas, Wasser, Abwasser 67 (1987) 111–122.

Giger, W., Brunner, P. H., Marcomini, A. and Siegrist, H.: Behaviour of LAS in sewageand sludge treatment plant. Tenside Surf. Det. 26 (2) (1989) 95–100.

Giolando, S., Rapaport, R. A., Larson, R. J. and Federle, T. W.: Environmental fate andeffects of DEEDMAC: a new rapidly biodegradable cationic surfactant for use infabric softeners. Chemosphere 30 (1995) 1076–1083.

Gledhill, W. E.: Review of the anaerobic biodegradation test methodology: relevanceto the environmental anaerobic conditions and detergent environmental data– A literature review for the US Soap & Detergent Association (1995).

González Mazo, E., Lara Martín, P. A. and Gómez Parra, A.: Determination and dis-tribution profiles of Alkyl Ethoxysulfates and Linear Alkylbenzene Sulfonates incoastal marine sediments from the bay of Cádiz (Southwest of Spain). Envir-onmental Toxicology and Chemistry, Vol. 24 (9) (2005) 2196–2202.

Guang-Guo Ying, Williams, B. and Kookana, R.: Environmental fate of alkylphenolsand alkylphenol ethoxulates. Environ. Itern. 28 (2002) 215–226.

Haggensen, F., Mogensen, A. S., Angelidaki, I. and Ahring, B. K.: Anaerobic treat-ment of sludge: focusing on reduction of LAS concentration in sludge. WaterScience and Technology, 46 (10, Sludge Management: Regulation, Treatment,Utilisation and Disposal), (2002) 159–165.

Harrits, S. M. and Hanson, R. S.: Stratification of aerobic methane-oxidizing organ-isms in Lake Mendota, Madison, Wisconsin. Limnology and Oceanography 25(1980) 412–421.

Heijnen, J. J., Mulder, A., Weltevrede, R., Hols, J. and Van Leeuwen, H. L. J. M.: Large-scale anaerobic-aerobic treatment of complex industrial wastewater using bio-film reactors. Water Science Technol. 23 (7–9) (1991) 1427–1436.

Heinze, J. and Britton, L.: Anaerobic biodegradation: environmental relevance, Pro-ceedings 3rd World Conference on detergents: global perspectives (ed.A. Cahn), p. 235, AOCS Press, Champaign, IL (1994).

HERA. Human and Environmental Risk Assessment of Linear Alkylbenzene Sulfo-nate. Available from <www.heraproject.com> (2004).

HMSO. The Assessment of Biodegradability in Anaerobic Digesting Sludge 1988.Methods for the Examination of Waters and Associated Materials. Her Majesty’sStationery Office: London, 1989.

Holt, M., Matthijs, E. and Waters, J.: The concentrations and fate of LAS in sludgeamended soils, Water Research, 23 (6) (1989) 749–759.

Holt, M. S. and Bernstein, S. L.: Linear alkylbenzenes in sewage sludges and sludgeamended soils. Wat. Res.; 26 (1992) 613–624.

Holt, M. S., Mitchell, G. C. and Watkinson, R. J.: The environmental chemistry, fateand effects of non-ionic surfactants. In: the Handbook of EnvironmentalChemistry , Vol. 3, Part F. Anthropogenic Compounds, Detergents. (ed. N. T. deOude), 89–144, Springer Verlag, Berlin, 1992.

Hon-Nami, H. and Hanya, T.: Linear alkylbenzene sulfonates in river, estuary andbay water. J. Jap. Limnol 41 (1–5) (1980).

Howarth, R. W. and Teal, J. M.: Sulfate reduction in a New England salt marsh. Lim-nology and Oceanography 24 (1979) 999–1013.

Huber, M., Meyer, U. and Rys, P.: Biodegradation mechanisms of linear alcholethoxylates under anaerobic conditions. Environ. Sci. Technol. 34 (2000)1737–41.

ISO 11734: Water quality – Evaluation of the ultimate anaerobic biodegradability oforganic compounds in digested sludge – Method by measurement of thebiogas production (1995).

ISO/DIS 14853: Evaluation of ultimate anaerobic biodegradation of plastic materi-als in an aqueous system – Method by analysis of carbon conversion to carbondioxide and methane (1999).

ISWA (International Solid Waste Association) Handling, treatment and disposal of ofsludge in Europe. Situation Report 1 ISWA Working Group on Sewage sludgeand and Water Works, 94 pages, Copenhagen, 1995.

Jensen, J.: Fate and effects of linear alkylbenzene sulphonates (LAS) in the terres-trial environment – a review. Sci. Tot. Environ. 226 (2–3) (1999) 93–111.

Jensen, J., Løkke, H., Holmstrup, M., Krogh, P. H. and Elsgaard, L.: Effects and riskassessment of linear alkylbenzene sulfonates in agricultural soil. 5. Probabilisticrisk assessment of linear alkylbenzene sulfonates in sludge-amended soils.Environ. Toxicol. Chem. 20 (8) (2001) 1690–1697.

Jensen, J. and Jepsen, S. E.: The production, use and quality of sewage sludge inDenmark. Waste Management 25 (3) (2005) 239–247.

Jensen, J., Smith, S. R., Krogh, P. H., Versteeg, D. J. and Temara, A.: Risk assessmentof LAS in agricultual soils revisited II: species sensitivity distribution and riskquotients. Chemosphere, in press (2007).

Jobst, H.: Chlorphenole und Nonylphenole in Klärschlämmen Teil 1: Vorkommen inKlärschlämmen westdeutscher Kläranlagen aus den jahren 1987 bis 1989.Acta hydrochim hydrobiol. 23 (1995) 20–25.

Jones, J. G.: Activities of aerobic and anaerobic bacteria in lake sediments and theireffect on the water column. In Sediment Microbiology, eds. D. B. Nedwell andC. M. Brown, 107–145. Academic Press: London, 1982.

Jørgensen, B. B.: Bacterial sulfate reduction within reduced microniches of oxidisedmarine sediments. Marine Biology 41 (1977a) 7–17.

Jørgensen, B. B.: The sulfur cycle of a coastal marine sediment (Limfjorden, Den-mark). Limnology and Oceanography 22 (1977b) 814–832.

Jørgensen, B. B.: Mineralization of organic matter in the sea bed – the role of sulfatereduction. Nature 296 (1982) 643–645.

Kaspar, H. F. and Tiedje, J. M.: Anaerobic bacteria and processes. In Methods of SoilAnalysis, Part 2. Chemical and Microbiological Properties, 989–1009. ASA-SSSA: Madison, 1982.

Kayser, R.: Upgrading the wastewater treatment plant of the city of Oldenburg. Wat.Sci. Tech. 29 (1994) 89–95.

Klein, Megauhey: The fate of detergents in septic tank systems and oxidationponds, Univ. of California, Report No. 64 (1), Jan. (1964).

Kloepper-Sams, P., Torfs, F., Feijtel, T. and Gooch, J.: Effects assessments for surfac-tants in sludge-amended soils: a literature review and perspectives for TRA,The Science of the Tot. Environment, 185 (1996) 171–185.

Klotz, H.: Alcohol ethoxylates (AE) in sewage treatment sludges. Interim results of aGerman ring test. Analytica Conference 98, April 21–24, München.

Krogh, P. H.: Holmstrup and Jensen J. (1997). NERI report N. 69 (1998) 16.Krogh, K. A., Vejrup, K. V., Mogensen, B. and Halling-Sorensen, B.: Liquid chroma-

tography mass spectrometry method to determine alcohol ethoxylates and al-kylamine ethoxylates in soil interstitial water, ground water and surface watersamples. Journal Chromatogr., A. 957 (2002) 45–57.

Kroiss, H., Schweighofer, P., Frey, W. and Matsché, N.: Nitrification inhibition – asource identification method for combined municipal and/or industrial waste-water treatment plants. Wat. Sci. Tech. 26 (1992) 1135–1146.

Küchler, T.: Behaviour of surfactants and their influence on the mobility of organicmicropollutants in sandy soils with weak sorption capacities, Fraunhofer Insti-tute, Thesis, Potsdam, 1995.

Kuhn, E. P., Zeyer, J., Eicher, P. and Schwarzenbach, R. P.: Anaerobic degradation ofalkylated benzenes in denitrifying aquifer columns. Appl. Environ. Microbiol. 54(1988) 490–496.

Kujawa, M., Schnaak, W. and Kuechler, T.: Occurrence of organic pollutants in sew-age sludge and influence of surfactants on their mobility in amended soils.Fraunhofer Internal Report (1996).

Kunkel, E.: Umweltanalytik von Tensiden. Tenside Surf. Det. 24 (1987) 280–285.Küster, E. and Niese, G.: Dumping of refuse and sludges. In Biotechnology Vol. 8.

Microbial Degradations, ed. W. Schönborn, VCH, (1986) 349–362.Lara Martín, P. A., Gómez Parra, A. and González Mazo, E.: Pressurized liquid ex-

traction followed by liquid chromatography-mass spectrometry for the deter-mination of mayor surfactants in marine sediments. International Journal ofEnvironmental Analytical Chemistry, Vol. 85 (4–5) April (2005) 293–303.

Lara-Martin, P. A., Gomez-Parra, A. and Gonzalez-Mazo, E.: Simultaneous extractionand determination of anionic surfactants in waters and sediments. Journal ofChromatography A , 1114 (2) (2006) 205–210.

Lara-Martin, P. A., Petrovic, M., Gomez-Parra, A., Barcelo, D. and González Mazo, E.:Presence of surfactants and their degradation intermediates in sediment coresand grabs from the Cadiz Bay area. Environmental Pollution 144 (2006) 483–491.

Lara-Martin, P. A., Gomez-Parra, A., Köchling, T., Sanz, J. L., Amils, R. and Gonzalez-Mazo, E.: Anaerobic degradation of linear alkylbenzene sulfonates in coastalmarine sediments, Environmental Science & Technology, 41 (10) (2007)3573–3579.



Lara-Martin, P. A., Gomez-Parra, A., Köchling, T., Sanz, J. L., Amils, R. and Gonzalez-Mazo, E.: Monitoring the primary biodegradation of linear alkylbenzene sulfo-nates and their coproducts in anoxic sediments using liquid chromatography-mass spectrometry, Environmental Science & Technology, 41 (10) (2007)3580–3586.

Larson, R. J., Federle, T. W., Shimp, R. J. and Ventullo, R. M.: Behaviour of LAS in soilinfiltration and ground water, Tenside Surf. Det. 26 (2) (1989) 116.

Latham, M. J.: The animal as an environment. In Microbial Ecology: A ConceptualApproach, eds. J. M. Lynch and N. J. Poole, 115–137. Blackwell Scientific Pub-lications: Oxford, 1979.

Laue, H., Denger, K. and Cook, A. M.: Taurine reduction in anaerobic respiration ofBilophila wadsworthia RZATAU, Appl. Environ. Microbiol., in press (1997).

Lee, H. B. and Peart, T. E.: Determination of 4-nonylphenol in effluents and sludgefrom sewage treatment plants, 67 (1995) 1976–1980.

León, V. M., Gonzalez-Mazo, E., Forja, J. M. and Gomez-Parra, A.: Identification oflinear alkylbenzene sulfonate biodegradation intermediates in anoxic marinecoastal sediments. Proceedings of the 5th World Surfactant Congress. ComiteEuropéen des Agents de Surface et Leurs Intermédiaires Organiques. Milano,Italy, 2000.

León, V., González Mazo, E., Forja Pajares, J. M. and Gómez Parra, A.: Vertical dis-tribution profiles of Linear Alkylbenzene Sulfonates and their long-chain inter-mediate degradation products in coastal marine sediments. EnvironmentalToxicology and Chemistry, Vol. 20 (10) (2001) 2171–2178.

Lie, T. J., Pitta, T., Leadbetter, E. R., Godchaux III, W. and Leadbetter, J. R.: Sulfonates:novel electron acceptors in anaerobic respiration, Arch Microbiol., 166 (1996)204–210.

López Serrano, I., Ferrer, J., Cassani, G., Ptavettoni, S. and Varo, P.: Study of LAS andsoap in marine sediments. Proceedings of the 32th annual CED meeting,(2001) 51–65.

Lowe, S. E., Jain, M. K. and Zeikus, J. G.: Biology, ecology, and biotechnological ap-plications of anaerobic bacteria adapted to environmental stresses in tem-perature, pH, salinity, or substrates. Microbiological Reviews 57 (1993) 451–509.

Madsen, T., Boyd, H. B., Nylén, D., Petersen, A. R., Petersen, G. I. and Flemming S.:Environmental and Health Assessment of Substances in Household Deter-gents and Cosmetic Detergent Products. Environmental Project No. 615, Mil-joeprojekt, CETOX (2001).

Madsen, T. and Rasmussen, H. B.: Evaluation of a method for determining theanaerobic biodegradability of surfactants. Report No. 38 VKI Water Quality In-stitute, Horsholm, Denmark, 1994.

Madsen, T., Rasmussen, H. B. and Nilsson, L.: Anaerobic Biodegradation Potentialsin Digested Sludge, Chemosphere 31 (1995) 4243–4258.

Madsen, T., Petersen, G., Seiero, C. and Torslov, J.: Biodegradability and AquaticToxicity of Glycoside Surfactants and a Nonionic Alcohol Ethoxylate; JAOCS, 73(1996) 929–933.

Madsen, T., Kristensen, P., Samsoe-Petersen, L., Torslov, J. and Rasmussen, J. O.:Application of sludge on farmland – quality objectives, level of contaminationand environmental risk assessment. Proc. Specialty conference on manage-ment and fate of toxic organics in sludge applied to land. Copenhagen, Den-mark, 30 April – 2 May 1997.

Mara, D. D. and Pearson, H.: Artificial freshwater environment: Waste stabilisationponds. In Biotechnology Vol. 8. Microbial Degradations, ed. W. Schönborn,(1986) 177–206.

Marcomini, A. and Giger, W.: Simultaneous determination of linear alkylbenzenesulfonates, alkylphenol polyethoxylates, and nonylphenol by high performanceliquid chromatography. Anal. Chem. 59 (1987) 1709 – 1715.

Marcomini, A., Cape, P. D., Lichtensteiger, Th., Brunner, P. H. and Giger, W.: Behav-iour of aromatic surfactants and PCBs in sludge- treated soil and landfills, J.Environ. Qual. 18 (1989) 523–528.

Marin, M., Pivotti, L., Campesan, G., Turchettto, M. and Tallandini, L.: Effects and fateof sediment sorbed linear alkylbenzene sulfonate (LAS) on the bivalve molluscMitilus galloprovincialis Lmk. Water Research 28 (1994) 85–90.

Matsché, N.: Kombinierter Einsatz von biologischer und chemischer P-Elimination.Gewässerschutz, Wasser, Abwasser 98 (1987) 13–33.

McAvoy, D. C., Eckhoff, W. S. and Rapaport, R. A.: Fate of Linear Alkylbenzene Sul-fonate in the environment. Environ. Tox. and Chem. 12 (1993) 977–987.

McAvoy, D. C., White, C. E., Moore, B. L. and Rapaport, R. A.: Chemical fate andtransport in a domestic septic system: sorption and transport of anionic andcationic surfactants. Env. Tox. Chem. 13 (1994) 213–221.

McEvoy, J. and Giger, W.: Accumulation of linear alkyl benzene sulfonates in sewagesludges. Naturwissenschaften 72 (1985) 429–431.

Matthijs, E. and De Henau, H.: Determination of LAS in aqueous samples, sedi-ments sludges and soils using HPLC. Tenside Surf. Det. 24 (1987) 4 (1987)193–199.

Matthijs, E., Debaere, G., Itrich, N., Masscheleyn, P., Rottiers, A. and Stalmans, M.:TW Federle, The fate of detergent surfactants in sewer systems, Wat. Sci.Tech. 31 (1995) 321–328.

Matthijs, E., Burford, M. D., Cassani, G., Comber, M., Eadsforth, C., Haas, P.,Klotz, H., Spilker, R., Waldhoff, H. and Wingen, H.-P.: Determination of AlcoholEthoxylate Components in Sewage Sludge. Tenside Surf. Det. 41 (3) (2004)113–120.

Metzner, G. and Lemmer, H.: Semi-Continuous Tests for Simulation of MunicipalSludge Digestion to Evaluate Anaerobic Processes. International SymposiumEnvironmental Biotechnology, Part II. Publisher: Technological 3 Institute, Ost-ende, 1997.

Metzner, G.: Elimination von Dodecylbenzolsulfonsäure und Methylglycindiessig-säure in anaerobem Milieu. Tenside Surf. Det. 35 (1) (1998) 46–51.

Metzner, G.: Ausgewählte aerobe und anaerobe Abbautestverfahren für Tenside.Münchener Beiträge zur Abwasser-, Fischerei- und Flussbiologie 54 (2001)263.

Mieure, J. P., Waters, J., Holt, M. S. and Matthijs, E.: Terrestrial safety assessment oflinear alkylbenzene sulfonate. Chemosphere 21 (1–2) (1990) 251–262.

Miranda, L. A. S., Henriques, J. A. P. and Monteggia, L. O.: Performance of UASband DAEB reactors in the anaerobic digestion of synthetic wastewater con-taining sodium oleate and sodium stereate. Water Science & TechnologyVol. 54 (2) (2006) 127–133.

Mix-Spagl, K.: Untersuchungen zum Umweltverhalten von Seifen. In: Umweltverträg-lichkeit von Wasch- und Reinigungsmitteln (Münchener Beiträge zur Abwasser-,Fischerei- und Flussbiologie, Vol. 44) R. Oldenburg Verlag, München, 1990.

Mogensen, A. S. and Ahring, B. K.: Formation of metabolites during biodegradationof linear alkylbenzene sulfonate in an upflow anaerobic sludge bed reactorunder thermophilic conditions. Biotechnol. Bioengineer. 77 (2002) 483–488.

Mogensen, A. S., Haagen, F. and Ahring, B. K.: Anaerobic degradation of Linear Al-kylbenzene Sulfonate. Environmental Toxycology and Chemistry, Vol. 22 (4)(2003) 706–711.

Mogensen, A. S., Hagensen, F. and Ahring, B. K.: Anaerobic biodegradation of linearalkylbenzene sulfonate. Environ. Toxicol. Chem. 22 (2003) 706–711.

Moreno, A., Ferrer, J. and Berna, J. L.: Biodegradability of LAS in a sewer system,Tenside Surf. Det. 27 (5) (1990) 312–315.

Moreno, A., Ferrer, J., Ruiz, B. F., Prats, D., Vazquez, B. and Zarzo, D.: Monitoring ofLinear alkylbenzene Sulfonate removal in a lagoon treatment system. CESIOinternational surfactants congress, London, 1992.

Moreno, A. Bravo, J., Ferrer, J. and Bengoechea, C.: Soap determination in sewagesludge by high-performance liquid chromatography. JAOCS 70 (1993) 667–671.

Morrall, S. W., Dunphy, J. C., Cano, M. L., Evans, A., McAvoy, D. C., Price, B. P. andEckhoff, W. S.: Removal and environmental exposure of alcohol ethoxylates inUS sewage treatment. Ecotoxicol. Environ. Saf.: 64 (2006) 3–13.

Mosey, F. E.: Anaerobic processes. In Ecological Aspects of Used-water Treatment.Vol. 2. Biological Activities and Treatment Processes, eds. C. R. Curds and H. A.Hawkes, 219–260. Academic Press, London, 1983.

Mosey, F. E.: Redox potentials in wastewater treatment. The Chemical Engineer,May 1985, 21–24.

Mountfort, D. O., Asher, R. A., Mays, E. L. and Tiedje, J. M.: Carbon and electron flowin mud and sandflat intertidal sediments at Delaware Inlet, Nelson, New Zeal-and. Applied and Environmental Microbiology, 39 (1980) 686–694.

Müller, M. T.: Anaerobic Biodegradation and Toxicity of Alcohol Ethoxylates. Thesis,ETH Zürich, 2000.

Nuck, B. A. and Federle, T. W.: A batch test for assessing the mineralisation of 14C-radiolabeled compounds under realistic anaerobic conditions. EnvironmentalScience and Technology 30 (1996) 3597–3603.

OECD. OECD Guideline 308 for the Testing of Chemicals: Aerobic and AnaerobicTransformation in Aquatic Sediment Systems. OECD, Paris, 2002.

OECD. Anaerobic Biodegradation of Organic Compounds in Digested Sludge –Method by Measurement of Gas Production. Draft New Guideline 311 (Octo-ber 2003 version). Organisation for Economic Cooperation and Development,Paris, 2003.

Osburn, Q. W.: Analytical methodology for linear alkylbenzene sulfonate (LAS) inwaters and wastes. J. Am. Oil. Chem. Soc. 63 (1986) 257–263.

Pagga, U. and Beimborn, D. B.: Anaerobic Biodegradation Test for Organic Com-pounds. Chemosphere 27 (1993) 1499.

Painter, H. A. and Zabel, T. F.: Review of the Environmental Safety of LAS. Report forECOSOL and SDIA, Water Research Centre, Medmenham, 1988.

Painter, H. A. and Zabel, T.: The behaviour of LAS in sewage treatment. Tenside Surf.Det. 26 (2) (1989) 108–115.

Painter, H. and Mosey, F.: The question of the anaerobic biodegradability of LAS.Proceedings of CESIO World Surfactant Congress, London (UK), 3 (1992) 34–43.

Painter, H. A.: Anionic surfactants. In The Handbook of Environmental ChemistryVol. 3 Part F, ed. O. Hutzinger, Springer-Verlag: Berlin, 1992.

Painter, H. A.: Detailed Review paper on Biodegradability Testing. OECD Paris, 1994.Petersen, G. I. and Andersen, T. T.: DHI report no. 10 2004. DHI, Denmark 2004.Petrovic, M., A., Fernandez-Alba, R., Borrull, F., Marce, R. M., Mazo, E. G. and Bar-

celò, D.: Occurrence and distribution of nonionic surfactants, their degradationproducts, and linear alkylbenzene sulfonates in coastal, waters and sedimentsin Spain. Environ. Toxycol. Chem. 2 (2002) 37–46.

Petzi, S.: Untersuchungen zum anaeroben Abbau von Seifen. Seifen-Öle-Fette-Wachse 115 (1989) 229–232.

Pittinger, C. A. and Kimerle, R. A.: Abstract #455, p.112. SETAC 11th Annual Meeting,Alexandria, VA (1991).

Prats, D., Ruiz, F., Vasquez, B. and Rodriguez-Pastor, M.: Removal of anionic andnonionic surfactants in a wastewater Treatment plant with anaerobic digestion.A comparative study. Wat. Res. 31 (8) (1997) 1925–1930.

Prats, D., Rodriguez, M., Varo, P., Moreno, A., Ferrer, J. and Berna, J. L.: Biodegrada-tion of soap in anaerobic digesters and on sludge amended soils. Wat.Res. 33 (5) (1999) 1331–1336.

Prats, D., Rodriguez, M., Varo, P., Muela, M., Llamas, J. M., Moreno, A., Ferrer, J. andBerna, J. L.: Elimination of xenobiotics during composting. A case study. LAS.Tenside Surf. Det. 36 (5) (1999) 294–298.

Press, F. and Siever, R.: Rivers: Currents, channels, and networks. In Earth, (1986)78–209.

Puchta, R., Krings, P. and Sandkühler, P.: A New Generation of Softeners. TensideSurf. Det. 30 (3) (1993) 186.

Randall, C. W., Barnard, J. M. and Stensel, H. D.: (Eds). Water Quality ManagementLibrary Vol 5: Design and retrofit of wastewater treatment plants for biologicalnutrient removal. Technomic Publishing Company, Lancaster, Pennsylvania,1992.

Rapaport, R. A. and Eckhoff, W. S.: Monitoring linear alkyl benzene sulfonate in theenvironment: 1973–1986. Environmental Toxicology and Chemistry 9 (1990)1245–1257.

Rapaport, R. A., Larson, R. J., McAvoy, D. C., Nielson, A. M. and Trehy, M.: The fate ofcommercial LAS in the environment. The CLER Review 1 (1995) 20–31.

Rehman, N., Melling, J., Van Duynhoven, J. and Roberts, D.: Fate of primary alkanesulfates (PAS) under anaerobic conditions – Effect of braching on biodegrad-ability. Poster presented at SETAC Europe, Lille, 2005.



Reiser, R., Toljander, H. and Giger, W.: Historic record of alkylbenzenesulfonates inrecent lake sediments: input changes and postburial fate of detergent derivedchemicals, Organic Geochemistry meeting, S. Sebastian, 1995.

Richards, F. R.: The Cariaco basin (Trench). Oceanography and Marine Biology An-nual Review 13 (1975) 11–67.

Richards, B. N.: The Microbiology of Terrestrial Ecosystems. Longman Scientific andTechnical: Harlow, 1987.

Rottiers, A., Boeije, G., Corstanje, R., Decraene, K., Feijtel, T., Matthijs, E. and Scho-wanek, D.: Adaptation of the CAS test system and synthetic sewage for bio-logical nutrient removal. Accepted for publication in Chemosphere (1998).

Rudd, J. W. M. and Hamilton, R. D.: Methane cycling in Lake 227 in perspective withsome components of the carbon and oxygen cycles. Archiv für HydrobiologieErgebnisse der Limnologie 12 (1979) 115–122.

Saez, M., Leon, V. M., Rubio, J. A., Gomez-Parra, A. and Gonzalez-Mazo, E.: A com-parative study of the adsorption of linear alkylbenzene sulfonates and theirbiodegradation intermediates on marine sediments and organisms. CienciasMarinas 29 (4) (2003) 397–403.

Salanitro, J. P., Wisneiwski, H. L., Byers, D. L., Neaville, C. C. and Schroeder, R. A.:Use of aerobic and anaerobic microcosms to assess BTEX biodegradation inaquifers. Ground water monitoring & Remediation 17 (3) (1997) 210–221.

Salanitro, J. P. and Diaz, L. A.: Anaerobic biodegradability testing of surfactants.Chemosphere 30 (1995) 813–830.

Sanderson, H., Price, B. B., Dyer, S. D., DeCarvalho, A. J., Robaugh, D., Waite, S. W.,Morrall, S. W., Nielsen, A. M., Cano, M. L. and Evans, K. A.: Occurrence andhazard screening of alkyl sulfates and alkyl ethoxysulfates in river sediments.Science of The Total Environment, 367 (1) (2006) 312–323.

Sanz, J. L., Rodriguez, M., Amils, R., Berna, J. L., Ferrer, J. and Moreno, A.: Anaerobicbiodegradation of LAS (Linear Alkylbenzene Sulfonate). Inhibition of themethanogenic process. La Rivista Italiana delle sostanze grasse, Vol LXXVI, Lu-glio/Agosto, 1999.

Sanz, J. L., Culubret, E., de Ferrer, J., Moreno, A. and Berna, J. L.: Anaerobic biode-gradation of linear alkylbenzene sulfonate (LAS) in upflow anaerobic sludgeblanket (UASB) reactors. Biodegradation 14 (2003) 57–64.

Sedlak, R. I. and Booman, K. A.: A Study of LAS and alcohol ethoxylate removal at amunicipal wastewater treatment plant. US SDA annual Convention, Boca Ra-ton, Fla (1986).

Sieburth, J. M.: Contrary habitats for redox-specific processes: Methanogenesis inoxic waters and oxidation in anoxic waters. In Microbes in the Sea, ed. M. A.Sleigh, 11–38. Ellis Horwood: Chichester, 1987.

Siegfried, M., Müller, M. T. and Baumann, U.: Anaerobic Degradation and Toxicity ofAlcohol Ethoxylates in Anaerobic Screening Test Systems; CESIO CongressBarcelona, 3 (1996) 61–275.

Shimp, R. J., Lapsins, E. V. and Ventullo, R. M.: Chemical fate and transport in do-mestic septic system : biodegradation of Linear Alkylbenzene Sulfonate (LAS)and Nitrilotriacetic Acid (NTA). Environmental Toxicology and Chemistry, 13 (2)(1994) 205–212.

Schlegel, H. G. and Jannasch, H. W.: Prokaryotes and their habitats. In The Prokar-yotes: A Handbook on Habitats, Isolation and Identification of Bacteria, Vol. 1,eds. M. D. Starr et al., 43–82. Springer-Verlag: New York, 1981.

Schöberl, P., Klotz, H. and Spilker, R.: (1996a). Alkylbenzolsulfonat (LAS)-Monitoring.Teil 3: LAS-Gehalte der Schwebstoffe verschiedener deutscher Fliessgewässer.Tenside Surf. Det. 33 (4) (1996) 329–335.

Schöberl, P. and Spilker, R.: (1996b). Alkylbenzolsulfonat (LAS)-Konzentrationen imLippe-Sediment eines Rhein-Altarmes. Tenside Surf. Det. 33 (5) (1996) 400–403.

Schowanek, D., Feijtel, T., Rottiers, A., Matthijs, E., Decraene, K., Boeije, G. and Ver-straete, W.: Development and validation of new continuous activated sludgetest systems with biological nutrient removal. Proc. IAWQ-NVA ConferenceAdvanced Wastewater Treatment, Amsterdam, September 23–25 (1996)431–433.

Schowanek, D., David, H., Francaviglia R., Hall, J., Kirchmann, H., Krogh, H. P.,Schraepen, N., Smith, S. and Wildemann, T.: Probabilistic Risk Assessment forLinear Alkylbenzene Sulfonate (LAS) in Sewage Sludge used on AgriculturalSoil. Regulatory Toxicology and Pharmacology 49 (2007) 245–259.

Shelton, D. R. and Tiedje, J. M.: General Method for Determining Anaerobic Biode-gradation Potential. Appl. Env. Microbiol. 47 (1984) 850–857.

Shokes, R. F., Trabant, P. K., Presley, B. J. and Reid, D. F.: Anoxic, hypersaline basin inthe northern Gulf of Mexico. Science 196 (1977) 1443–1446.

Smolenski, W. J. and Suflita, J. M.: Biodegradation of cresol isomers in anoxic aqui-fers. Applied and Environmental Microbiology 53 (1987) 710–716.

Steber, J. and Wierich, P.: The Anaerobic Degradation of Detergent Range Fatty Al-cohol Ethoxylates. Studies with 14C-Labelled Model Surfactants, Water Res. 21(1987) 661–667.

Steber, J., Gode, P. and Guhl, W.: The ecological evaluation of a group of importantdetergent surfactants – fatty alcohol sulfates. Soap/Cosmetics/Chemical Spe-cialities 64 (1988) 44–50.

Steber, J. and Wierich, P.: The environmental fate of fatty acid �-sulfomethyl esters:biodegradation studies with a carbon – 14 labelled model surfactant. Tenside,26 (1989) 406–411.

Steber, J.: Wie vollständig sind Tenside abbaubar?, Textilveredelung 26 (1991) 348–354.

Steber, J. and Birch, R. R.: Chemical structures and their biodegradability in theanaerobic environment. Final Report of EU Project Contract No. STEP-CT 91-0154 (1995a).

Steber, J., Herold, C.-P. and Limia, J. M.: Comparative evaluation of anaerobic bio-degradability of hydrocarbons and fatty derivatives currently used as drillingfluids. Chemosphere 31 (1995b) 3105–3118.

Steber, J., Guhl, W., Stelter, N. and Schröder, F. R.: (1995c). Alkyl polyglycosides –ecological evaluation of a new generation of nonionic surfactants, Tenside Surf.Det. 32 (6) (1995) 515–521.

Strotmann, U. J. and Eglsäer, H.: The toxicity of substituted phenols in the nitrifica-tion inhibition test and luminescent bacteria test. Ecotoxicol. Environ. Saf.,30 (3) (1995) 269–273.

Swanwick, J. D., Shurben, D. G. and Jackson, S.: A survey of the performance ofsewage sludge digesters in Great Britain. Journal of the Institute of Water Pol-lution Control 6 (1969) 639–661.

Sweetman, A. J.: Development and application of a multi-residue analytical methodfor the determination of n-alkanes, linear alkylbenzenes, polynuclear aromatichydrocarbons and 4-nonylphenol in digested sewage sludges. Wat. Res. 2!(1994) 343–353.

Tabor, C. F. and Barber, L. B.: Fate of linear alkylbenzene sulfonate in the Mississippiriver. Environ. Sci. Technol. 30 (1996) 161–171.

Takada, H. and Ishiwatari, R.: Linear alkylbenzenes in urban riverine environments inTokyo: distribution, source, and behaviour. Environ. Sci Technol. 21 (1987)875–883.

Takada, H., Ishiwatari, R. and Ogura, N.: Distribution of linear alkylbenzenes (LABs)and linear alkylbenzene sulfonates (LAS) in Tokyo Bay sediments. Estuarine,Coastal and Shelf Science 35 (1992) 141–156.

Temper, U., Pfeiffer, W. and Bischofsberger, W.: Stand und Entwicklungspotentialeder anaeroben Abwasserreinigung unter besonderer Berücksichtigung der Ver-hältnisse in der Bundesrepublik Deutschland. Berichte aus Wassergüte-wirtschaft und Gesundheitsingenieurwesen, Nr. 67, Technische UniversitätMünchen, 1986.

Tschui, M. and Brunner, P. H.: Die Bildung von 4-Nonylphenol aus 4-Nonylphenol-mono- und diethoxylat bei der Schlammfaulung, Vom Wasser, 65 (1985) 9–19.

US-EPA. Design manual Phosphorus Removal. EPA Technology Transfer Report EPA625/1-87/001, Center for Environmental Research Information, Cincinnati,OH 45268 (1987).

Vandepitte, V. and Debaere, G.: Biodegradation (aerobic/anaerobic) of dimethyldodecyl amine oxide. Procter & Gamble ETC, Poster SETAC Congress SanFrancisco, 1997.

Varo Galván, P., Prats Rico, D. and Rodriguez Pastor, M.: Anaerobic biodegradationof soaps in anaerobic screening test. Tenside Surf. Det. 39 (5) (2002) 160–163.

Verge, C., Moreno, A. and Berna, J. L.: (1997). Influence of water hardness on thebioavailability and toxicity of LAS. Chemosphere 44 (2001) 8.

Wagener, S. and Schink, B.: Anaerobic degradation of nonionic and anionic surfac-tants in enrichment cultures and fixed-bed reactors, Wat. Res. 21 (1987) 615–622.

Wahlberg, C., Renberg, L. and Wideqvist, U.: Determination of nonylphenol andnonylphenol ethoxylates as their pentafluorobenzoates in water, sewagesludge and biota. Chemosphere 20 (1990) 179–195.

Waters, J. and Feijtel, T. C. J.: AISE/CESIO Environmental surfactant monitoring pro-gramme; outcome of five national pilot studies on linear alkylbenzene sulfo-nate (LAS). Chemosphere 30 (1995) 1939–1956.

Zehnder, A. J. B. and Stumm, W.: Geochemistry and biogeochemistry of anaerobichabitats. In ‘Biology of Anaerobic Microorganisms’, ed. A. J. B. Zehnder, 1–35,Wiley-Liss. Ahel, M., Giger, W. (1985). Determination of alkylphenols and alkyl-phenol mono-and diethoxylates in environmental samples by high perform-ance liquid chromatography. Anal. Chem. 57 (1988) 1577–1583.

Glossary of Abbreviations

y AE : Alcohol Ethoxylatesy AES : Alcohol Ether Sulfatesy AISE : Association International de la Savonnerie,

de la D�tergence et des Produits d’Entretieny AOS : Alpha Olefin Sulfonatesy APG : Alkylpoliglucosidesy APEO : Alkyl Phenol Ethoxylatesy AS : Alcohol Sulfates, Alkyl Sulfatesy BABS : Branched Alkylbenzene Sulfonatey CEFIC : European Federation of Chemical Industriesy CESIO : Comit� Europ�en des Agents de Surface et

leurs Interm�diaires Organiquesy DIC : Dissolved Inorganic Carbony DM : Dry Mattery DTDMAC : Di Tallow Di Methyl Ammonium Chloridey ECETOC : European Centre of Ecotoxicology and Toxi-

cology of Chemicalsy EEA : European Environment Agencyy EPA : Environmental Protection Agencyy ERASM : Environmental Risk Assessment and Man-

agementy HAD : Haupt-Ausschuss Detergentieny HERA : Human and Environmental Risk Assess-

menty HPLC : High Performance Liquid Chromatographyy ILSI : International Life Sciences Institutey ISO : International Standarisation Officey LAS : Linear Alkylbenzene Sulfonates



y LeAF : Lettinga Associates Foundationy MES : Methyl Esther Sulfonatesy OECD : Organisation of Economic Cooperation and

Developmenty SAS : Secondary Alkane Sulfonatesy SCHER : Scientific Committee on Health and Envir-

onmental Risksy SPC : Sulpho Phenyl Carboxilatesy STP : Sewage Treatment Plant

y TMAC : Tri Methyl Alkyl ammonium Chloridey UASB : Up Flow Anaerobic Sludge Blanket reactor

Appendix

Overview of surfactant monitoring and mass balancing inanaerobic environmental compartments

Name Surfac-tant

Type Monitoringsite

Concentrationlevels

(mg/kg DM)

Environ-mental

conditions

Removalresults(%)

Resi-denceTime

Remarks References

Anionics

Sulfonates LAS Commercial Digestorsludges

2900–11900 Anaerobic M McEvoy & Giger(1985)

LAS Commercial Digestorsludges

4200 (1200) Anaerobic 20–30%

29 digestors/grabsamples

Giger et al., (1987)


9300–18800 Anaerobic M Grab samples; onedigestor

Holt & Bernstein(1992)


6660 Anaerobic 0% Sedlak et al.,(1986)


4660 (1540) Anaerobic M US five plants –49 grabs

Rapaport & Eckhoff(1990)


6000–9400 Anaerobic M Monitoring in Spain& Italy

Waters & Feijtel(1995)

LAS Commercial Riversediment

100–322 undefined M Osburn (1986)

LAS Commercial river SS 2–209 notapplicable

M Schöberl et al.,(1996)


0–3.3 Undefined M Monitoring cores ofLippe sedimentfrom 1939–1991

Schöberl & Spilker(1996)

LAS Commercial Anaer. pondwater

5.2–6.3 mg/l UndefinedEh

ca. 20% 20 to60 days

Over a period ofone year

Moreno et al.,(1994)

LAS Anaer. pondsedim.

43600 Moreno et al.,(1994)

LAS Tridecyl-benzene

Fresh waterpond

0.48 (spiked) Anaerobic 0% Lab study withpond sediment in-

oculum

Federle & Schwab(1992)


0.49–5.3 undefined M Monitoring in Spain& Italy

Waters & Feijtel(1995)


0.01–20 undefined M Monitoring in Mis-sisipi

Tabor & Barber(1996)


0.6–567 undefined M Monitoring in Tokyoarea

Takada & Ishiwatari(1987)


5.6 undefined M Marcomini & Giger(1987)

LAS Commercial Marinesediment

10 to 30 UndefinedEh

M Matthijs et al.,(1986)


11 to 30 UndefinedEh

M Hon-Nami & Hanya(1980)

LAS Commercial Landfill 245 (bottom)– 9160 (top)

UndefinedEh

M Landfill receivingdigested sludge

Marcomini et al.,(1989)


0.5 to4.74 mg/kg

M Monitoring of Bot-tom sediments inPo river (Italy)20 points

Cavalli et al.,(2000)


1.2–26.7 mg/kg in summer1.2–62.1 in

winter

M 5 points in Cadizbay and 2 points in

Barbate river

DelValls et al.,(2002)



Name Surfac-tant

Type Monitoringsite

Concentrationlevels

(mg/kg DM)

Environ-mental

conditions

Removalresults(%)

Resi-denceTime

Remarks References

Sulfonates LAS Commercial Marinesediment

0.5–1 mg/kgin coast

2–20 mg/kg inport

M Branched sulfo-nates were alsodetected in some

samples

Folke et al., (2003)


138 mg/kgclose todischarge

0.8–16.4 inother points

SPC in the range0.9–0.07 mg/kgwere also detected

Leon V. M. et al.,(2001)

LAS Commercial Marine sedin Elbeestuary

0.039–0.109 mg/kg

M Several locations inthe Elbe estury at

5 m depth

Bester et. al.,(2001)

LAS Commercial Sludgeamende

soil

<1 mg /kg20 mg/kg in

heavyamended soilfor 25 years

8 field sites in theNorth of Spain

Carlsen (2002)

LAS Commercial Marine sedin estuary

0.538–1.014 mg/kg

M Guadalete estuary(Es)

Lara-Martin et al.,(2006)

LAS Commercial Sludgeamended

soil

0.12–12.84 mg/kg

for LAS0.004–10.220

for SPC

M LAS and SPC in 10field sites, 10daysto 4 years aftersludge treatment

Eichhorn, Peteret al., (2005)

SAS Commercial Digestorsludges

648–738 Anaerobic 0% 40 days 10 day sampling Field et al., (1995)

Sulfates AES SepticTanks

Anaerobic 72–81% No environmentcharacterisation

Birch et al., (1992)

AS Digestor 66% to be completed

Soaps Soap Digestorsludges

18000–51900 Anaerobic 70–75% 26 days Sampling over sev-eral treatment

plants

Moreno et al.,(1993)

Nonionics

AlcoholEthoxylates

AE Commercial Digestor 2600 (400) Anaerobic 65% 26 days One day sampling Prats et al., (1997)

AE Commercial Digestor <500 typically Anaerobic M Monitoring pro-gramme in Ger-many (Tegewa)

Klotz (1998)

Alkylphe-nol Ethoxy-

lates

NP Digestorsludges

900 (600) Anaerobic M 29 digestors Giger et al., (1987)

NP Digestorsludges

450–2530 Anaerobic M Several digestors Giger et al., (1984)

NP Digestorsludges

1600 Anaerobic M Ahel & Giger(1985)

NP +NPEO

Digestorsludges

3–540 Anaerobic M Madsen et al.,(1997)

NP Digestorsludges

35–95 Anaerobic M Kujawa et al.,(1996)

NP Digestorsludges

545–1000 Anaerobic M Non Indicated Ahel et al., (1994)

NP Digestorsludges

130–400 Anaerobic M Kunkel (1987)

NP Digestorsludges

640–2200 Anaerobic M Brunner et al.,(1988)

NP Digestorsludges

137–470 Anaerobic M Lee & Peart (1995)

NP Digestorsludges

1200 Anaerobic M Marcomini & Giger(1987)

NP Digestorsludges

400–1200 Anaerobic M Wahlberg et al.,(1990)

NP Digestorsludges

20–350 Anaerobic M Chalaux et al.,(1994)



Name Surfac-tant

Type Monitoringsite

Concentrationlevels

(mg/kg DM)

Environ-mental

conditions

Removalresults(%)

Resi-denceTime

Remarks References

Alkylphe-nol Ethoxy-

lates

NP Digestorsludges

21–1193 Anaerobic M Jobst (1995)

NP Riversediment

0.29–6.73 undefined M Lee & Peart (1995)

NP Riversediment

NP River sedi-ment


NP Riversediment

0.02–0.04 undefined M Chalaux et al.,(1994)

NP Marine se-diment

0.006–0.07 undefined M Chalaux et al.,(1994)

NP Subsurface(aquifier)

0.2–0.96 lg/l(range 0.0–

33)

Mixed Eh/Aerobic

90% 2–4 h Soil biodegradationof NP

Ahel et al., (1996)

NP1EO Digestorsludges

5–40 Anaerobic M Kunkel (1987)


< 3 Anaerobic M Kunkel (1987)










20–190 Anaerobic M Wahlberg et al.,(1990)


1.0–50.0 Anaerobic M Wahlberg et al.,(1990)

NP1EO Riversediment


NP2EO Riversediment


NP1EO Subsurface(aquifier)

0.04–0.91 lg/l(range 0.0–

4.9)

Mixed Eh/Aerobic


Ahel et al., (1996)

NP2EO Subsurface(aquifier)

0.01–0.33 lg/l(range 0.00–

23)

Mixed Eh/Aerobic


Ahel et al., (1996)

NP1EC Subsurface(aquifier)

2.9–10.9 lg/l(range 0.00–

13.1)

Mixed Eh/Aerobic


Ahel et al., (1996)

Received: 07. 08. 2007Revised: 02.10. 2007

y Correspondence to

Dr. José Luis BernaPetresa GroupCampo de Las NacionesAvenida Partenon, 12–14E-28042 MadridSpainE-mail: [email protected]

The authors of this paper

José Luis Berna Tejero got his degree in chemistry at the University of Zaragoza(Spain) in 1969 and master in chemistry and technology of petroleum at the Uni-versidad Complutense of Madrid in 1971. He joined Petresa in 1971 as a researchchemist and he is presently Director of R&D of the Petresa group. He is the authorand co-author of numerous publications and participations in international con-gresses mainly in the field of surfactant raw materials development, environmentalchemistry and risk assessment of surfactants.

Giorgio Cassani studied chemistry and obtained his PhD at the University of Milan,Italy. From 1976 to 1982 he worked at the research centre Instituto Donegani(Montedison Group), first in the field of synthesis and identification of new “insectpheromones” of agricultural interest and later in the field of isolation/-identificationof microbial fermentation products (biotechnology group). In 1992 he moved toEniChem Augusta (now Sasol Italy S.p.A.) as Head of the Analytical Laboratory atthe research centre in Paderno Dugnano (Milano). His main experience now is inthe field of surfactant: analysis, environmental determination and biodegradationstudies. He is author/co-author of more than 40 publications.



Claus-Dierk Hager holds a German diploma in chemistry and obtained his PhD inmetal organic chemistry at the University of Dortmund in 1979. In 1979 he joinedUnilever as a process development manager in the edible fats and oils business. In1984 he moved to Hüls AG, Surfactants Business Unit, where he was responsiblefor the application of surfactants in detergents and cleaners. After the acquisition ofthe Hüls Surfactants Business by CONDEA Chemie GmbH and finally in 2001 bySASOL, Claus-D Hager is now working for SASOL Germany, where he is managingthe Product Safety and Regulatory Affairs Department as well as Market Research.

Naheed Rehman is an environmental team leader advising Unilever on all aspectsof environmental safety. She holds a PhD and conducts global environmental riskassessments for Unilever’s ingredients. She is responsible for delivery of Unilever’sSafety & Environmental Assurance Centre (SEAC’s) ecotoxicology work programme.Dr. Rehman leads and co-ordinates R&D projects for safety approval with specialistexpertise in assessing the effect of chemical structure on the biodegradability ofUnilever’s ingredients (aerobic and anaerobic). She has published several reportsand scientific papers in peer reviewed journals.

Ignacio López Serrano obtained his MSc in Chemistry by the University of Granadain 1997. He got a Master in Engineering and Environmental Management by theEscuela de Organización Industrial in 1999. At the end of 1999 he joined Petresaworking within the R+D division. His area of research comprises of the study ofraw materials, surfactant derivatives and detergency as well as the environmentalbehaviour of surfactants. Currently he is responsible for the Customer Service andEnvironmental Laboratory of Petresa.

Diederik Schowanek is a Bio-Engineer by training and obtained his PhD in Environ-mental Engineering in 1991 from Gent University (Belgium). He is currently princi-pal scientist in the Product Safety and Regulatory Affairs Department of Procter &Gamble’s Brussels Innovation Centre (Belgium). Dr. Schowanek is involved in theplanning and management of environmental (research) activities of P&G. He hasbuilt a recognised expertise in domains like Risk Assessment of chemicals and LifeCycle Assessment. Dr. Schowanek is also actively involved in professional organisa-tions such as B-IWA (Belgian Committtee of the International Water Association),EWA (European Water Association), and the Royal Flemish Engineering Society

(KVIV- Section ‘Environmental Technology’). With P&G he is also involved in variousEU-sponsored environmental research projects. Dr. Schowanek has published over40 scientific papers in international journals and handbooks.

Josef Steber holds a diploma in biology and obtained his PhD in microbiology/bio-chemistry at the University of Munich in 1976. He joined the Department of Ecologyat Henkel in 1978. Until his retirement end 2006 he was in charge of this depart-ment and has been member of numerous international working groups dealingwith the environmental safety assessment of chemicals.

Klaus Taeger was born in 1954. He studied microbiology at the University of Göttin-gen and obtained his PhD in Wuppertal in 1986. Until 1988 his main subject wasthe bioremediation of soil. Afterwards he has worked in the departments of Emis-sion Control and Ecology as well as Product Safety at BASF AG. An important part ofhis work is to investigate the biodegradability of detergent ingredients.

Thorsten Wind holds a German Diploma of Biology and an American Bachelor ofScience. He finished his PhD-work in 1997 at the Max-Planck-Institute of TerrestrialMicrobiology in Marburg/Germany. After two years in the field of international pro-ject management in the diagnostic industries Thorsten Wind started 1999 as a labmanager of the ecological laboratories at Henkel KGaA- Düsseldorf/Germany. He iscurrently working as a manager of Environmental Product Safety Assessment in theDept. of Corporate SHE & Product Safety responsible for the detergent sub-division.His special expertise is in exposure analysis, environmental monitoring and model-ling and biodegradation.

You will find the article and additional material by enter-ing the document number TS100351 on our website atwww.tsdjournal.com

