Review

Refinement of biodegradation tests methodologies and the proposedutility of new microbial ecology techniques

Agnieszka Kowalczyk a,n, Timothy James Martin b, Oliver Richard Price c,Jason Richard Snape d, Roger Albert van Egmond c, Christopher James Finnegan c,Hendrik Schäfer a, Russell James Davenport b, Gary Douglas Bending a

a School of Life Sciences, University of Warwick, Coventry CV4 7AL, United Kingdomb School of Civil Engineering and Geosciences, Newcastle University, Newcastle upon Tyne NE1 7RU, United Kingdomc Unilever, Safety & Environmental Assurance Centre, Colworth Science Park, Sharnbrook MK441LQ, United Kingdomd AstraZeneca, Mereside, Alderley Park, Macclesfield SK10 4TF, United Kingdom

a r t i c l e i n f o

Article history:Received 30 April 2014Received in revised form22 September 2014Accepted 23 September 2014

Keywords:OECD testsBiodegradationPersistenceChemical risk assessmentMicrobial ecologyOmics

a b s t r a c t

Society's reliance upon chemicals over the last few decades has led to their increased production,application and release into the environment. Determination of chemical persistence is crucial for riskassessment and management of chemicals. Current established OECD biodegradation guidelines enabletesting of chemicals under laboratory conditions but with an incomplete consideration of factors thatcan impact on chemical persistence in the environment. The suite of OECD biodegradation tests do notcharacterise microbial inoculum and often provide little insight into pathways of degradation. Thepresent review considers limitations with the current OECD biodegradation tests and highlights novelscientific approaches to chemical fate studies. We demonstrate how the incorporation of molecularmicrobial ecology methods (i.e., ‘omics’) may improve the underlying mechanistic understanding ofbiodegradation processes, and enable better extrapolation of data from laboratory based test systems tothe relevant environment, which would potentially improve chemical risk assessment and decisionmaking. We outline future challenges for relevant stakeholders to modernise OECD biodegradation testsand put the ‘bio’ back into biodegradation.

& 2014 Elsevier Inc. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102. OECD biodegradation tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10

2.1. Historical aspects and principal design of OECD test. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102.2. Overview of current tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 112.3. Limitations of current tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

2.3.1. Ready biodegradability tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 132.3.2. Enhanced and modified screening tests within REACH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 162.3.3. Inherent biodegradability tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16

3. Environmental realism of tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 173.1. Effect of environmental factors on biodegradability assessment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

4. Microbial ecology in biodegradation testing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 174.1. Advantages and disadvantages of ‘omics’ . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19

5. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19

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Ecotoxicology and Environmental Safety

http://dx.doi.org/10.1016/j.ecoenv.2014.09.0210147-6513/& 2014 Elsevier Inc. All rights reserved.

n Corresponding author. Present address: SC Johnson, Frimley Green, Frimley, Camberley, Surrey GU16 7AJ, United Kingdom. Tel.: þ44 7758572996.E-mail address: AKowalcz@scj.com (A. Kowalczyk).

Ecotoxicology and Environmental Safety 111 (2015) 9–22

Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

1. Introduction

Ever since Silent Spring (Carson, 1962) there has been wide-spread public concern about the use of chemicals and theirpossible impact on the environment. Silent Spring facilitated theban of the pesticide DDT for agricultural use in 1972 in the UnitedStates (US EPA, 1975; Grier, 1982) and the onset of greaterregulatory and public interest in the environmental consequencesof chemicals. The beginning of the chemical regulatory era withinthe EU can most likely be traced back to the demise of the‘principle of microbial infallibility’ which was based on the beliefthat, given the opportunity and favourable conditions, any organicchemical would biodegrade (Painter, 1974). Ironically, syntheticsurfactants produced in the 1950s under the concept of ‘betterliving through chemistry’ (Copley et al., 2012) were found toproduce unsightly foaming in conventional wastewater treatmentplants, leading to the realisation that they were not beingdegraded during treatment. The 1950s and 1960s heralded anera of biodegradability testing development centralised predomi-nantly around the aerobic biodegradability of synthetic detergents(Allred et al., 1964; Bunch and Chambers, 1967).

As our awareness of environmental issues associated with theuse of chemicals has increased over the past 40–50 years, so has thescientific understanding and regulatory requirements. However, inparallel, the global chemical industry has grown rapidly since 1970.Global chemical output (produced and shipped) was valued atUS$171 billion in 1970. By 2010, it had grown to $4.12 trillion(Davies, 2009). Pesticide use has increased worldwide by 36 fold inthe last 45 years (1960–2005) (Zhang et al., 2011). Global use ofpharmaceuticals led to 112% increase in prescription drug salesrecorded between 2000 and 2008 (FDA, 2013) and the global use ofhome and personal care products has increased by 232% and 750%between 1998 and 2013, respectively (Euromonitor, 2013). Thesetrends are likely to continue for the foreseeable future, as growth inemerging markets in South America, Africa and Asia continues.

Various regulations have been developed to assess the adverseimpacts of existing and new chemicals in the environment. Hazarddata are generated to assess the persistence, bioaccumulation andtoxicity of chemicals. Exposure models and emission scenarioshave been developed to predict the distribution and transport ofchemicals in the environment. A key component of both hazarddetermination (persistence) and environmental risk assessment isthe accurate estimation of the biodegradation of a chemical in theenvironment.

The degree to which any individual chemical will partition andpersist in the environment is governed by a number of intrinsicand extrinsic properties/factors. These include solubility/hydro-phobicity, octanol/water partition coefficient (Kow), soil organiccarbon/water partition coefficient (Koc), dissociation constant(pKa), as well as characteristics of the environmental compartment(e.g., organic carbon content, soil/sediment particle size and viablenumber of bacteria). Microbial biodegradation of chemicals is themajor process that affects chemical persistence in the environ-ment (Copley, 2009). Although several OECD biodegradation testguidelines have been established to determine chemical fate underdifferent environmental scenarios, the prediction of chemicalfate is still challenging (e.g. biodegradation studies conductedat environmentally relevant concentrations, laboratory to fieldextrapolation).

Since information on the fate and behaviour of chemicals isrequired to make an assessment of environmental exposure, it iscrucial that chemical fate is determined in an accurate manner andmore important to be predictive of degradation rates in theenvironment. However, so far the chemical industry, regulatorsand academia have encountered difficulties with extrapolatingdata from standardised laboratory tests into the environment dueto discrepancies between test conditions and the complexity ofenvironmental conditions. Studying microbial interactions andtheir functions within inocula helps to link processes associatedwith chemical biodegradation at the community level. Suchknowledge can be essential to better understand chemical biode-gradation in OECD tests and in the environment. Application ofmicrobial ecology methods in existing test systems can be used tooptimise new test guidelines. They may also have the potential inhelping to develop strategies to improve the reproducibility ofOECD tests.

In this review we provide an overview of the current testmethods used to assess chemical biodegradation and discuss theirlimitations. A review of the current OECD tests is essential tooutline the key issues. e.g., microbial inocula, chemical concentra-tion, biodegradation pass levels. It will enable stakeholders (aca-demia, industry and government) to understand the majorconcerns regarding current test results and identify knowledgegaps/questions. New methods can then be proposed to addressthese questions. Finally, we provide a summary of technologicaldevelopments which could improve the capacity of biodegrada-tion tests to provide more reliable predictions of biodegradation inthe complexity of real environments.

2. OECD biodegradation tests

2.1. Historical aspects and principal design of OECD test

A range of methods for investigating biodegradation processeshave been developed to predict the fate of chemicals in theenvironment. Most efforts have focused on the fate of chemicalsin the aquatic environment, especially in wastewater treatmentprocesses (Reuschenbach et al., 2003). Testing biodegradabilityunder laboratory conditions aims to obtain a reliable prediction ofthe likelihood of the biodegradability of chemicals in the environ-ment (Pagga, 1997). Over 30 years ago, in 1981, the OECD firstpublished its guidelines for testing the biodegradation of chemicals.There have been several amendments and additions since, includ-ing updated methods for assessing ready biodegradability in 1992(OECD, 1992) and introduction of the CO2 headspace test (OECD 310,2006). In 1990, a classification in accordance with the OECD wasproposed (OECD, 2005). Three groups of tests were defined:(1) ready biodegradability (or screening), (2) inherent biodegrad-ability and (3) simulation (Lapertot and Pulgarin, 2006). A list ofreference chemicals for use as positive and negative controls instandardized biodegradability tests (Table 1) has been proposed byregulators and industry with an agreed set of properties andcharacterised set of biodegradability behaviour, which cover a rangeof environmental persistence and non-persistence (Comber andHolt, 2010). These chemicals group into bins (Table 1), which alignwith OECD tiered testing (Fig. 1) and show the relationships

A. Kowalczyk et al. / Ecotoxicology and Environmental Safety 111 (2015) 9–2210

between the screening and higher tier tests (Comber and Holt,2010).

2.2. Overview of current tests

The choice of a test should depend on the testing purpose, andunder REACH, the testing is graded according to production leveland properties of a chemical (ECHA, 2012). Ready biodegradability

tests (RBTs) are considered a stringent first tier and indicate if achemical is rapidly degradable or not. Seven methods (Table 2)permit the screening of chemicals for ready biodegradability in anaerobic aqueous medium and they are based on the removal oforganic chemicals measured as dissolved organic carbon (DOC)(OECD 301 A, OECD 301 E), the production of the catabolic end-product carbon dioxide (OECD 301 B), the determination of thebiochemical oxygen demand (BOD) (OECD 301C, OECD 301 D,

Fig. 1. Relationship between screening and higher tier biodegradability tests and the bins.Adapted from Comber and Holt (2010)

Table 1Reference chemicals and their classification into bins.Adapted from Comber and Holt (2010)

Bin Description Half-life(days)

Chemical example

1 Reference chemicals which would normally pass a RBT or modified RBT test o15 Aniline, sodium benzoate, phenol2 Reference chemicals that would normally pass an enhanced screening biodegradability test but currently fail

any other tests16–40 4-Chloroaniline, 4-fluorophenol, 1,2,3-

trimethylbenzene3 Reference chemicals that would normally fail any biodegradability screening test whether modified RBT or

enhanced screening biodegradability test41–60 o-Terphenyl, cyclodecane,

dibutylphenol4 Reference chemicals that should never pass a modified RBT or an enhanced biodegradability screening test 460 Hexachlorobenzene, benzo(a)pyrene,

hexachlorohexane

RBT – ready biodegradability tests.

A. Kowalczyk et al. / Ecotoxicology and Environmental Safety 111 (2015) 9–22 11

OECD 301 F) (Reuschenbach et al., 2003; OECD, 1993) and bymeasuring the inorganic carbon (IC) produced in the test bottles(OECD 310, 2006).

The second tier tests are inherent tests where inherent biodegrad-ability can be measured by specific analysis (primary biodegradation)

or by non-specific analysis (ultimate biodegradation). Inherent biode-gradability tests include Modified SCAS Test (OECD 302 A, 1981),Zahn-Wellens/EMPA Test (OECD 302 B, 1992) and Modified MITI Test(II) (OECD 302C, 1981) (Table 2). The test procedure offers a higherchance of detecting biodegradation compared to tests for ready

Table 2Overview of OECD biodegradation tests.

Biodegradationtest

OECDguideline

Pass level Incubationconditions

Chemicalconcentra-tion

Inoculum source Inoculum size Testduration

Readybiodegrad-ability tests(screeningtests)

OECD 301 A 70% DOC removal Aerobic 10–40 mgDOC/l

Activated sludge, sewageeffluents, surface waters,soils or mixture of these

r 30 mg/l settled sewage;r100 ml effluent/l; approx.107–108 cells/l

28 Days

OECD 301 B 60% ThCO2 10–20 mgDOC/l

OECD 301 C 60% ThOD 100 mg/l Fresh samples from sewagetreatment works, industrialWWTPs, soils, lakes, seas,mixed thoroughly together

30 mg/l settled sewage;approx. 107–108 cells/l

OECD 301 D 2–10 mg/l or Derived from secondaryeffluent of WWTP orlaboratory-scale unitpredominantly domesticsewage, alternativelysurface water e.g., river, lake

r5 ml effluent/l; approx.104–106 cells/l5–10 mg

ThOD/l

OECD 301 E 70% DOC removal 10–40 mgDOC/l

Derived from secondaryeffluent of WWTP orlaboratory-scale unitpredominantly domesticsewage

0.5 ml effluent/l; approx..105 cells/l

OECD 301 F 60% ThOD 100 mg/l or50–100 mgThOD/l

Activated sludge, sewageeffluents, surface waters,soils or mixture of these

r30 mg/l settledsewage; r100 mleffluent/l; approx.. 107–108 cells/l

OECD 310 60% ThIC 20–40 mg C/l Activated sludge, sewageeffluents, surface waters,soils or mixture of these

4–30 mg SS/l or 10% v/vsecondary effluent

Inherent(potential)biodegrad-ability tests

OECD 302 A 4 20% ThBOD, ThDOCremoval or ThCOD (primarybiodegradation; 420% ThBOD, ThDOC removal orThCOD (ultimatebiodegradation)

Aerobic 2 –10 mg/l Mixed settled sludges aftertwo weeks aeration period

A high concentration ofaerobic micro-organisms

Not defined

OECD 302 B 50–400 mgDOC/l

Activated sludge 0.2–1.0 g dry matter/l

OECD 302C 30 mg/kg Activated sludge 100 mg/kg

Simulationtests

OECD 303 A Estimation of half-lives formaterials exhibiting first-order degradation patterns.In the absence of first-orderkinetics degradation timesfor 50%, (DT50) and 90%(DT90) may be reported

Aerobic/anaerobic

41–100 μg/l Activated sludge 2.5 g/l dry matter; 2–10 ml/leffluent

120 Days

OECD 303 B Airborne inoculation,settled sewage

1 ml/l of settled sewage

OECD 307 Representative soil; a sandyloam or silty loam or loamysand with a pH of 5.5–8.0,organic carbon content of0.5–2.5%

Microbial biomass of atleast 1% of total organiccarbon

OECD 308 Sediments from samplingsites selected based on thehistory of possibleagricultural, industrial ordomestic inputs to thecatchment and the watersupstream

Not defined

OECD 309 Surface water fromsampling sites selectedbased on the history ofpossible agricultural,industrial or domesticinputs

Not defined

OECD 314(A–E)

Raw wastewater, activatedsludge, anaerobic digestersludge, treated effluent-surface water mix,untreated effluent-surfacewater mix,

Not defined

DOC – dissolved organic carbon; ThBOD – theoretical biochemical oxygen demand; ThCOD – theoretical chemical oxygen demand; ThCO2 – theoretical carbon dioxide;ThDOC – theoretical dissolved organic carbon; ThIC – theoretical inorganic carbon; ThOD – theoretical oxygen demand; WWTP – wastewater treatment plant.

A. Kowalczyk et al. / Ecotoxicology and Environmental Safety 111 (2015) 9–2212

biodegradability and therefore, if an inherent test is negative thiscould indicate a potential for environmental persistence (Comber andHolt, 2010). The highest tier tests are simulation tests (Table 2) whichaim at assessing the rate and extent of biodegradation in a laboratorysystem designed to represent either the aerobic treatment stage of awastewater treatment plant (WWTP) or environmental compart-ments, such as fresh or marine surface water.

The majority of OECD tests are conducted under environmen-tally unrealistic conditions e.g., the chemicals are tested at highconcentrations that are unlikely to occur in the environment.In addition, the test volumes vary from test to test, there is lackof consideration of the inoculum quantity or quality, and protocolsregarding inoculum preparation prior to biodegradability testingalso vary. Furthermore the tests are conducted under standardizedincubation conditions, which do not reflect highly variable envir-onmental conditions, for example seasonality but do allow easeand conservatism in regulatory risk assessment. Other limitationsinvolve lack of consistency in the pass levels and test duration.

2.3. Limitations of current tests

In order to provide reliable degradation data for predictingchemical fate, biodegradation tests should aim to mimic realenvironmental exposure and represent small-scale equivalents ofreal environments. It should be noted that biodegradation testsonly provide some of the information needed to determineenvironmental fate of chemicals. The physico-chemical conditionsand the nature of the microbial populations present in theenvironment are difficult to achieve in biodegradation tests andmay never be truly realised, particularly given the probability ofadequately obtaining relatively rare organisms representatively insmall samples.

2.3.1. Ready biodegradability testsRBTs are primary screening tests which have historically been

the central foundation for assessing the biodegradation of chemi-cals in regulatory frameworks. This can be attributed to their easeof use, relatively straightforward interpretation and the amount ofdata collated since their introduction (Aronson et al., 2006). Thesehighly prescribed, standardised, and conservative regulatory testsmeasure the relative biodegradability of chemicals. They typicallymeasure ultimate biodegradation (the biodegradation of thechemical to carbon dioxide, biomass, water and other inorganicchemicals) but can be used to measure the primary biodegrada-tion. A positive result in a RBT is defined as a 60% reduction intheoretical oxygen demand (ThOD) or theoretical CO2 evolution(ThCO2) or a 70% reduction in DOC within the 10-day window(beginning at 10% degradation) of a 28-day test. Fulfillment ofthese criteria can be considered as indicative of rapid and ultimatebiodegradation in the environment (OECD, 2005). However, RBTstypically exhibit high levels of variation including: inter-replicatevariation, inter-test variation, inter-facility variation and temporalvariation (Nyholm et al., 1984; Painter, 1995). As a result, theyproduce a large number of test fails, many of which may beconsidered false negatives, whereby a chemical fails a test notnecessarily due to poor biodegradability but rather because it failsthe test design. Test failure (i.e. insufficient degradation of the testcompound within the 10-day window of a 28-day test), may beindicative of the persistent nature of the test compound, but it canalso occur as a result of a number of different factors: there mayonly be partial degradation (less than 60% ThOD/ThCO2 or lessthan 70% DOC removal); there may be sufficient degradation butnot within the confines of the 10-day window of a 28-day test; thechemical may have been introduced at an initial concentrationtoxic to the microbial inocula; the low inocula concentrations in

some tests may reduce the probability of incorporating competentdegraders; the artificial mineral media used in standard tests maynot be suitable for supporting the growth and establishment ofspecific degraders. False negatives are thought to account for20–80% of RBT fails (ECETOC, 2007) and have been recognised asa limitation since the introduction of the tests (ECETOC, 1983). It isdifficult to accurately qualify the occurrence of false negatives withdata as this information is not readily released by industry,however the variation within and between tests has previouslybeen reported in academic literature.

The high-levels of variation exhibited in RBTs as a result of thepreviously mentioned factors were demonstrated by Gerike andFischer (1979) who conducted a thorough assessment of pre-OECD301 RBTs (recommendations from which were incorporated into theOECD 301 series) and reported significant differences in biodegrada-tion outcome dependent upon the selected method. In the case of4-nitrophenol, biodegradation ranged from 0% to 100% dependentupon the chosen method (Gerike and Fischer, 1979). This wassupported by Thouand et al. (1995) who developed a probabilisticmodel which calculated a probability of degradation for 4-nitrophenolranging from 0% to 99% depending upon the method selected.Nyholm et al. (1992) reported that screening tests underestimatedbiodegradability for 4-chloroaniline, desmethyl-methyl parathionand in some cases 4-nitrophenol, where extensive lag phasesresulted in false negative test fails. These compounds provided apositive biodegradation outcome in simulation tests (Nyholm et al.,1992). Žgajnar Gotvajn and Zagorc-Končan (2003) reported a test failfor diethylene glycol in a closed bottle test (OECD 301 D) but a testpass in a respirometric test (OECD 301 F). In comparing two differentrespirometric test systems, Reuschenbach et al. (2003) reported thepotential for the same test to return conflicting degradation datadependent upon the test system applied. A thorough literaturereview conducted by Comber and Holt (2010), reported errantbehaviour in standard tests and significant variations in reportedhalf-lives in non-standard tests for a range of compounds including4-nitrophenol, 4-chloraniline and 4-fluorophenol.

In contrast with some of the fixed test variables such as testduration, the biological criteria suffer from a lack of adequatecharacterisation. Variations in the source, concentration and pre-treatment of the inoculum have previously been shown to impactupon the diversity of test inocula and their ability to degradechemicals (Goodhead et al., 2014), which can add to the source ofvariability and large number of false negatives associated withRBTs (Vázquez-Rodríguez et al., 2011). The potential to produce morereliable and reproducible tests by increasing research efforts sur-rounding the test inoculum has been identified in workshopssponsored by the European Centre for Ecotoxicology and Toxicologyof Chemicals (ECETOC) (ECETOC, 2003, 2007).

2.3.1.1. Inoculum source, concentration and preparation. A centralassumption in all freshwater biodegradability tests is that arandomly selected sample from an environmental compartment willexhibit the diversity required to provide the requisite range ofmicroorganisms for assessing the biodegradation potential of aparticular chemical in any given environmental system. Notwithstan-ding a historical understanding of the importance microbes have inbiodegradation (Sweeney and Foote, 1964), the outcome is presentlydependent upon the potential of an arbitrarily selected mixedcommunity of relatively low concentration to degrade a chemicalunder standardised conditions. Typically, biodegradation is viewedwith respect to the intrinsic properties of the chemical and rarely fromthe perspective of the bacteria, namely the probability of encounteringspecific degraders of the test chemical within the test inocula.

OECD guidelines stipulate that inocula may be sourced from:activated sludge, sewage effluent, surface waters, soils, a mixture

A. Kowalczyk et al. / Ecotoxicology and Environmental Safety 111 (2015) 9–22 13

of the aforementioned, rivers, a 10 site composite sample orderived from secondary effluent at varying concentrations. There-fore, OECD screening tests can involve different inocula celldensities which can affect chemical biodegradation. For example,in activated sludge and river water the number of specificdegraders may vary from 3.7�106 to 2.5�104 cells/l, respectively(Thouand et al., 1995, 2011). The modified French Associationfor Standardization (AFNOR) method specified use of 5�102 cells/lin test assays and the modified Sturm test suggested10�2�102 cfu/l (Painter, 1995). It was reported by Thouandet al. (1996) that only 105 degrader cells/l are needed for a 99.9%chance of biodegradation of 4-nitrophenol. Hence, number of cellscan vary by four orders of magnitude between tests. In theircurrent form, screening tests can be considered as much a test ofthe inoculum as a test of chemical biodegradability. The variationin inoculum concentration also results in significant differences infood to microorganism ratios (F:M), which may impact upondegradation kinetics: where the substrate concentration is highand the microbial biomass low, there is more potential for growthlinked biodegradation and higher rates of CO2 evolution; as thesubstrate concentration decreases and microbial biomass levelsincrease, biodegradation kinetics change with typically lessgrowth and thus less CO2 evolution which can impact upon theoutcome of the test (Blok and Booy, 1984; Blok et al., 1985).

Adaptation is defined as a change in the microbial communitythat increases the rate of transformation of a test chemical as aresult of a prior exposure to the test chemical and plays a key rolein chemical biodegradation (Spain and van Veld, 1983). Pretreat-ment or preadaptation of the inocula is not permitted under OECDguidelines. However, preconditioning of the inocula is permitted,including filtration, centrifugation, settling and decantation andaeration. These treatments have been shown to normaliseobserved degradation potential characteristics but reduce theefficacy of the inocula to degrade even readily biodegradable testchemicals (Vázquez-Rodríguez et al., 2007). This is probably due tothe removal of selected members of the microbial communities,reducing overall diversity and leading to the exclusion of rarespecific degrader organisms (Goodhead et al., 2014)

Degradability and persistence are often thought of as proper-ties intrinsic to the chemical in question. In fact they are complexproperties influenced by a number of factors, of which thepropensity of a given chemical structure to biodegrade is onlyone. Understanding the relationship between microbial diversityand biodegradation outcome, the occurrence and abundance ofcompetent degraders and the bacterial variation between inoculawill ultimately facilitate the design of more robust screeningstudies and enable better predictions of biodegradation. This canbe achieved by inoculum characterisation (e.g. cell density, micro-bial community composition, functional diversity and biodegrada-tion potential) and inoculum standardisation (e.g. filtration orcentrifugation, in order to obtain a standardized requiredinoculum size).

2.3.1.2. Test chemical concentration. There is a growing focus on theproblem of extrapolation of biodegradability data obtained fromstandard laboratory tests to conditions observed in the field. OECDscreening and inherent tests assess rates of biodegradation atchemical concentrations generally far in excess of those likely tobe found in natural ecosystems (Lapertot and Pulgarin, 2006).Simulation tests for biodegradation of chemicals in surface watersystems provide single scenario information about biodegradationrates, and adaptation behaviour at environmentally realisticconcentrations (at least below 100 μg/l and normally at or below10 μg/l) (Comber and Holt, 2010). In most screening tests10–400 mg of carbon per litre is applied, while the concentration

of chemicals in WWTPs or in the environment are usually in therange of ng/l to μg/l (Ahtiainen et al., 2003; Kolpin et al., 2002).Extrapolations using data from such tests at high concentrationsmay not reflect what takes place at low concentrations (Lapertotand Pulgarin, 2006). Toräng et al. (2003) demonstrated that directuse of the biodegradation rate data obtained for mecoprop and 2,4-dichlorophenoxyacetic acid at high (4100 μg/l) concentration in anaerobic aquifer, grossly overestimated the actual degradation rates atenvironmentally-relevant concentrations, which were in the order of0.1 μg/l. However, Kern et al. (2010) demonstrated that laboratorybiodegradation rates of pharmaceuticals and biocides and formationof their transformation products reflected their removal rates duringactivated sludge treatment.

Concentration of the chemical is a significant factor affecting itsbehaviour in the environment and therefore, its susceptibility tomicrobial attack (Efroymson and Alexander, 1995), which isdefined by the Michaelis–Menten kinetics of enzymatic reactions(Chou and Talaly, 1977). Chemicals at low concentrations do notserve as primary substrates but are more likely to be degraded assecondary (non-growth) substrates concurrently with a variety ofnaturally occurring carbonaceous chemicals (Berg and Nyholm,1996). Presence of organic substrates and nutrients in thoseenvironments might lead to preferential degradation orco-metabolism of other chemicals and lower biodegradation ratesof the primary chemical (de Lipthay et al., 2007; Markiewicz et al.,2011). Furthermore, presence of some chemicals may cause aninhibitory effect on microbial populations. Stasinakis et al. (2008)reported the toxic effect of triclosan (TCS) and nonylphenol (NP)on activated sludge heterotrophic and autotrophic microorgan-isms, and a possible risk for deterioration of nitrification inactivated sludge systems due to the presence of TCS at unrealis-tically high concentrations.

Chemical concentration affects biodegradation kinetics. Forexample, Berg and Nyholm (1996) found that the percentage ofaniline, 4-chloroaniline and pentachlorophenol removed by una-dapted sludge had a tendency to be higher when the test chemicalwas dosed at trace concentrations (10 μg/l), than at standard(high) concentrations (20 mg DOC/l). In contrast, degradation ofother chemicals (halobenzoate, nitroacetic acid, 2,4-dichlorophe-nol and isopropyl N-phenylcarbamate) can be limited at lowconcentrations (Pitoi et al., 2011). The biodegradation of differentchemicals at low concentration varies, possibly due to the differentconcentrations needed for the specific enzymes and chemicalbiodegradation pathways to be induced in bacterial cells (Grady,1984).

Extents and rates of microbial adaptation, important for biode-gradation behaviour, can also be greatly influenced by the chemi-cal concentration. Low concentrations can be expected to increasethe period needed for adaptation in comparison with higherconcentrations (Berg and Nyholm, 1996). Subsequently, manyorganic chemicals persist in the environment. The reason for thiscould be that time is needed for the population of degraders tobecome sufficiently large or active to bring about chemicaldegradation or for microorganisms to synthesise and/or activatethe required enzymes for biodegradation (Grady, 1984). In somestudies, concentration thresholds have been reported below whichno adaptation could be detected. Efroymson and Alexander (1995)reported a lack of phenanthrene mineralisation in a mixed cultureat concentrations of 0.6–20 μg/l.

2.3.1.3. Inoculum. According to Thouand et al. (1996) at least threeparameters must be controlled when running a biodegradationtest. These are the physico-chemical conditions (pH, temperature,oxygen, concentration of mineral elements), the concentration ofthe chemical to be tested, and the concentration of the bacterial

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inoculum. While it is possible to reproduce the first two parameters inbiodegradation tests, the inoculum represents a significantchallenge for the reproducibility of test results. Microbial inoculaused in OECD biodegradation tests are one of the most poorlycontrolled variables. High variability in the biodegradability ofchemicals in OECD tests is potentially due to the characteristics ofthe inocula (presence or absence of biodegraders, adaptation to thechemical, cell density, and presence of protozoa) which control bothlag phase and biodegradation kinetics, which may lead toproblematic interpretation (Thouand et al., 1995). Unfortunately,bacterial density estimation methods are rarely applied and so farthere are no practical methods/protocols to measure inoculumdiversity in industrial laboratories (van Ginkel et al., 1995;Thouand et al., 2011). Therefore, there is a need for othertechniques that should be used in parallel biodegradation testingto determine not only the overall size of inoculum but also itsquality (number of specific degraders) and the diversity offunctional communities involved in biodegradation processes.

Biodegradability tests may involve inocula from different ori-gins such as: river water, sea water, activated sludge and soil(Thouand et al., 2011). Since inoculum diversity is related to itsorigin it should be taken into account while running tests. Forexample, Forney et al. (2001) demonstrated that the diversity ofseveral activated sludges differed worldwide. This could reflectdifferences in environmental factors such as temperature, humid-ity and characteristics of wastewaters received by the WWTPs, butmay also represent differences in the global biogeographicaldistribution of microbial taxa (Prosser et al., 2007). Moreover,specific degraders for non-readily degradable chemicals are morelikely to be found in the sediment of rivers and coastal areas thanin the water, whereas samples from pristine areas contain verysmall numbers of specific degraders (Thouand et al., 2011).Franklin et al. (2000) reported higher cell numbers and microbialdiversity in contaminated zones of ground water than in pristineareas. Additionally, pristine aquifers are carbon-limited environ-ments, with bacterial populations in a state of starvation andreduced activity (Goldscheider et al., 2006). An inoculum withhigh species diversity has a greater chance to contain degradingspecies and therefore, to degrade test chemicals in comparisonwith less diverse inoculum. Margesin et al. (2003) reported thatcontaminated environments harbour a wide range of degradingmicroorganisms, which play a crucial role in the bioremediationprocesses.

2.3.1.3.1. Biofilm as inoculum. There are several reasons whybiofilms should be considered in chemical fate studies (screeningand simulation studies). They are heterogenic, comprising com-plex communities of algae, cyanobacteria and bacteria embeddedin a polysaccharide matrix attached to substrates such as rocks,sediments and aquatic plant surfaces (Vinten et al., 2011). Biofilmaccumulation in riverbeds or on suspended particles contributeslargely to the removal of chemicals from the water, and thus thebiofilm enhances self-purification of surface water (Muia et al.,2003; Edwards et al., 1990).

In rivers and streams, biofilms are the first to interact withdissolved chemicals such as nutrients, organic matter, and chemi-cals. Biofilms integrate the effects of environmental conditionsover extended periods of time, mainly because of their rapidgrowth, species richness, and the physiological variety of theorganisms of which they are formed. They have been used forroutine monitoring and could be useful as ‘early warning systems’of ecosystem disturbances (Sabater et al., 2007). Brümmer et al.(2000) observed stability of river biofilm community structure inspite of significant variation in chemical and biological parametersduring the year. Araya et al. (2003) demonstrated the importantcontribution of natural river biofilm communities to the riverdie-away test conducted under laboratory conditions. Biofilms

had a significant participation in the biodegradation capacity ofdiazinon-contaminated river ecosystems (Tien et al., 2011). Naturalriver biofilms developed on ceramic disks showed 99.9% removalof diazinon at an initial concentration of 5 mg/l, over 26 days, inthe laboratory tests. Since biodegradation of many groups ofchemicals requires relatively rare species to be present in thetested inoculum (Thouand et al., 2011) biofilms have potential asan inoculum since they consist of complex bacterial communities,which represent natural microbial populations.

Characterisation of microbial diversity within biofilms and itschanges after exposure to organic chemicals is crucial to under-stand the effects of chemicals on aquatic ecosystem. For instance,Yergeau et al. (2012) determined the active community composi-tion within river biofilms and subtle changes in microbial com-munities were detected after short-term exposure to lowconcentrations (0.5–1.0 μg/l) of the antibiotics erythromycin, sul-famethoxazole and sulfamethazine.

2.3.1.4. Test volume. If biodegradation does occur, it is generallypresumed that the bacteria responsible are rare organisms with aminimum abundance of one competent degrader in the initialinoculum. This presumption inherently concerns degradation byspecific degraders capable of metabolising the chemical inquestion and may not be true if, for example, co-metabolism ofthe chemical as a non-growth substrate occurs first, yielding anintermediate more amenable to common degradation pathways(Horvath, 1972). In using low density inocula, independent of inoculatest volumes, the probability of encountering rare organisms, even inreplicate samples, is reduced and their inclusion is a matter ofchance (Thouand et al., 1995; Ingerslev and Nyholm, 2000; Ingerslevet al., 2000). RBTs are performed typically at low volumes, whichincreases the likelihood of variation in microbial compositionbetween tests, ultimately impacting upon the reproducibility of thebiodegradation outcome within and between facilities. Thisphenomenon refers to the decreasing probability of encounteringspecific degraders of the test chemical within the test inocula, withdecreasing test volume, or decreasing inocula concentration(Nyholm et al., 1992).

2.3.1.5. Test duration. Even from early development there appearto have been concerns over the robustness of RBTs with respect tothe arbitrary test duration assigned during development (Painter,1995). RBTs proceed for 28 days. In addition to passing the requiredextent of biodegradation in this period, a ‘readily biodegradable’chemical must also degrade at least as fast as the minimum raterequired to satisfy the 10-day window (see below). Painter (1995)argued that the test duration might have been adopted due toconcerns regarding reductions in analytical sensitivity beyond 28days, potentially impacting upon data reliability. Advancements inanalytical sensitivity since RBT introduction, however, would almostcertainly negate this concern. The 28 day test duration has beenutilised since the early development of biodegradability testing inthe 1960s and it is possible that this time period has been retainedbecause: it has been used previously; there is a plethora of datatesting at this time point; and until recently (ECETOC, 2003, 2007) ascientific case for change has not been presented.

The ‘10-day window’ opens at 10% degradation, a time pointthought to indicate the cessation of the lag phase. Painter (1995)suggested the window may exist in reference to the 7–10 daystypically required for primary standard degradation in the originalOECD guidelines for assessing biodegradability of anionic surfac-tants. Following their introduction, little research was conductedregarding 10-day windows and few studies utilised them duringtheir analysis at the time, making it difficult to assess theirnecessity.

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The 28 day test duration and 10-day window restrict the timeframe within which to observe the growth of specific competentdegraders. The relatively high chemical concentration combinedwith the low microbial biomass levels stipulated in some testsselects for degraders which proliferate rapidly and/or have asufficiently dense population in the original sample (Thouandet al., 2011). Extending the test duration beyond 28 days inpersistency assessments takes into account the extensive lagphases exhibited during the testing of some compounds andmakes allowances for poorly water soluble substances (ECHA,2012). The need to extend screening tests beyond 28 days toimprove chemical assessments and characterisation of persistencyhas been recognised within REACH and proposed as an enhancedmethod (ECHA, 2008).

A chemical exhibiting a slower rate of degradation, for exampledegrading in 60 days compared to 28 days, may pose a higher riskto human and environmental health, considering the impact ofchronic, low-level chemical exposure, the effects of which are notfully understood (Kümmerer, 2010). However, to date there is noevidence to support the idea that a chemical which undergoescomplete mineralisation in a longer test is more likely to pose amore serious threat to the environment than a chemical whichfulfils current RBT pass criteria.

RBTs are not designed to be representative of the environment,they are in fact deliberately not environmentally realistic, howeverthe arbitrary test durations, particularly the 10-day window, donot provide a full account of degradation in the environment andultimately lead to more expensive, intensive testing.

2.3.1.6. Pass threshold. A chemical is classified as readily biodegrad-able if one of the following requirements is fulfilled: a 60%reduction in ThOD (theoretical oxygen demand) or production ofThCO2 (theoretical CO2), or a 70% reduction in DOC. A ring testperformed shortly after the initial introduction of the 301 seriessuggested that whilst the high DOC removal rate was justified, therewas evidence to support a reduction in ThOD and ThCO2 from 60%to reflect that part of the test chemical carbon will be utilised innew cell synthesis. The study found in some cases that a ThOD ofless than 60% correlated with a DOC removal of greater than 90%(Painter and King, 1985). The ring test led to some modificationswhich were implemented in the modified OECD guidelines (OECD,1992) but the initial pass thresholds were maintained, presumablyto maintain high stringency of the tests (Weytjens et al., 1994).Arguably a more relevant biodegradation measure would be thehalf-life (time point at which 50% degradation is reached). Recently,environmental concerns have developed at a faster rate than theseinternational standards, with a regulatory shift from screening outchemicals which undergo rapid mineralisation in all environmentsto identifying chemicals which persist in the environment, a qualitytypically assigned based on biodegradation half-lives (ECETOC,2007). Although screening tests are not effective as screens forpersistence and were not originally designed to assign half-lives forchemicals, they are increasingly being used to do so in light of thisshift (Aronson et al., 2006).

2.3.2. Enhanced and modified screening tests within REACHFor the aforementioned reasons, RBTs are highly variable and

often produce false negative results. They are also unsuitable, intheir current guise, for making assessments of persistency. Persis-tency is difficult to measure as it is not an intrinsic property of thechemical per se. It is typically inferred based on a continuedconcentration in the environment and a lack of observed degrada-tion data in laboratory studies. One of the aims within REACH is toeffectively assess persistence for use in prioritisation, and to thisend a number of enhancements and modifications to existing

screening studies have been identified (ECHA, 2008). Modifiedready biodegradability tests incorporate modifications to RBTsaimed to address issues such as toxicity and bio-availability.Proposed modifications to screening tests include the use ofbiosurfactants or humic acid for testing chemicals with poor watersolubility and testing at lower substrate concentrations to circum-vent toxicity issues. Providing the other RBT criteria are fulfilled,they can still be used to assess ready biodegradability, whereasenhanced biodegradation screening tests should be used to assistin persistency assessments and not as an indication of readybiodegradability. The proposed enhancements were selected toimprove the environmental relevance of biodegradability assess-ments and reduce the volume of simulation studies required(ECHA, 2008). The ultimate aim is the design of a test which liessomewhere between the more stringent primary screening studies(RBTs) and the more expensive (both fiscally and with respect toresources) higher tier tests (Comber and Holt, 2010). Recom-mended enhancements focus on (1) increasing study lengthbeyond 28 days, (2) use of semi-continuous test systems,(3) increased vessel size and inoculum density, with environmen-tal samples not previously exposed to the test chemical and(4) running two ready tests in series (ECETOC, 2007; Comberand Holt, 2010). The modifications and enhancements proposedunder REACH address a number of the commonly discussedreasons for RBT variability and false negatives. The ultimate aimof the amendments is to provide a more environmentally relevantinoculum, which is particularly addressed by those strategieswhich ensure a more representative diversity (e.g. increasinginoculum concentration and test volume), which will allow betterpredictions of persistency (ECHA, 2008).

2.3.3. Inherent biodegradability testsThe predominant focus of this critique, and of proposed

enhancements to existing studies, is on screening tests andparticularly RBTs. In addition to their reduced complexity, incomparison with inherent biodegradability studies and simulationtests, and their ease of interpretation, RBTs are also those tests thatmost testing facilities will predominantly perform. The majority ofexperimental data available for ‘tested’ chemicals is in the form ofRBT assessments. According to the eChem portal (OECD, 2013), ajoint collaboration between the OECD and the European ChemicalsAgency (ECHA) which collates chemical data from multiplesources and databases, more than four times as many RBTs havebeen performed as inherent tests when assessing chemical biode-gradability (4604 and 1013 respectively).

Inherent biodegradability tests are less stringent than RBTs andas such the test parameters are more relaxed. They are designed toidentify chemicals which, given the opportunity, show an inherentpotential for degradation. Despite this difference, many of thedrawbacks associated with RBTs are still applicable to inherenttests, particularly with respect to the inoculum and the testduration. The inoculum is still subject to preparation methodswhich are thought to reduce diversity, although not as much as inRBTs (Thouand et al., 2011). Inocula concentrations remain vari-able between tests with no characterisation of the inoculum orconcentration in the majority of inherent tests. The positive aspectwith regards to the inoculum in the OECD 302 A guideline is theincreased environmental relevance gained in some tests by usinginocula directly from source, rather than using unrealistically lowmicrobial concentrations (in RBT/screening tests) in dilutedinocula or artificially amended systems.

The 302 B Zahn-Wellens inherent test is subject to the arbitrarilyassigned 28 day test duration. However, the 302 A and 302C aregiven varying test durations dependent upon experimental observa-tions which provides a more reflective representation of behaviour in

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the environment. There also remain concerns, for inherent tests andRBTs alike, over issues such as the varying manner of endpointanalysis and the suitability of the artificial mineral media forsupporting the establishment and growth of competent degradercommunities.

3. Environmental realism of tests

There are OECD tests simulating freshwater aquatic (OECD 309,2004; OECD 314, 2008) and terrestrial compartments (OECD 307,2002), based on static, semi-continuous, or continuous operationunder aerobic or anaerobic conditions with different test media.The test conditions are standardized as far as possible and attemptto reflect the environmental reality of biodegradation processes.Although, simulation tests are applicable for use in higher tierbiodegradation test, their complex test design, long test durationand high costs, make them unsuitable as routine tests (Merrettig-Bruns and Jelen, 2009).

3.1. Effect of environmental factors on biodegradability assessment

Screening tests often underestimate the potential of chemicaldegradation in environmental systems (Guhl and Steber, 2006).This could be due to a variety of factors, which differ between laband natural environments for instance, pH, light/dark cycle,oxygen, flow rate of water body and river bed form. This sectionis not comprehensive and Table 3 presents only some examples ofenvironmental factors that affect chemical biodegradation in theenvironment, and which are not fully covered by OECD tests thatare conducted under standardized laboratory conditions. CurrentOECD tests are not considered as environmentally realistic. Forexample, light conditions are excluded from biodegradation teststo avoid the growth of algae, which may affect the biodegradationof tested chemicals and/or make the interpretation of CO2 datadifficult due to fixation of the CO2. Interestingly, some authorshave reported that light enhances biodegradation of chemicals

(Borde et al., 2003; Li et al., 2005) due to growth of phototrophicorganisms capable of chemical biodegradation (Roldán et al., 1998;Lima et al., 2003; Thomas and Hand, 2011, Davies et al., 2013).The environmental factors can have both positive and negativeeffects on degradation rate. For instance, Eriksson et al. (2002)reported that lower temperature severely limited the biodegrada-tion of PAHs under aerobic conditions. However, Lartiges andGarrigues (1995) observed different effect of temperature onbiodegradation kinetics of pesticides in different types of water(sea water, river water, filtered river water). For example, theyreported small influence of temperature on triazine biodegrada-tion contrary to the carbamate carbaryl. Temperature is also animportant environmental variable which affects bacterial growthand biomass (White et al., 1991). Temperature has a stronginfluence on the period of acclimation of microorganisms and onthe rate of biodegradation (Kang and Kondo, 2002; Manzano et al.,1999) therefore, it may affect the outcome of biodegradation tests.pH may vary widely (from pH o3 to 49) within and betweenenvironmental compartments and may alter the toxicity of che-micals (Rutgers et al., 1998; Araoye, 2009) and affect the biode-gradation rate (Li et al., 2008). Furthermore, chemicals can bedifferentially mineralised in aerobic and/or anaerobic environ-mental compartments (Boyd et al., 1983; Battersby and Wilson,1989). In natural river systems attenuation of chemicals may beinfluenced by sediment dynamics and flow velocity which deter-mine contact of chemicals with bacteria on the sediment surface(Kunkel and Radke, 2008). The size and structure of biofilm is alsoaffected by flow shear stress (Lau and Liu, 1993; Battin et al., 2003).

4. Microbial ecology in biodegradation testing

Biodegradation tests play the key role in chemical persistenceassessment upon which environmental risk assessment is per-formed. Current OECD tests have been designed in the mannerthat does not consider microbial inocula and there is little under-standing of chemical breakdown pathways and how they relate to

Table 3Examples of environmental factors that affect chemical biodegradation.

Variables Test chemical Degradation testþresults Reference

pH 4-Nitrophenol Degradation rate affected by pH, optimum pH¼8–9 Li et al. (2008)Pentachlorophenol (PCP) Toxicity and degradation of PCP is affected by pH Rutgers et al. (1998)Hexachlorocyclohexane Optimum conditions were found for biodegradation at pH¼8 and T¼30 1C Siddique et al. (2002)

Temperature Bisphenol A Longer biodegradation at 20 1C than at 30 1C with half-lives from 4–7 days to2–6 days, respectively

Kang and Kondo (2002)

Nonylphenol polyethoxylate Biodegradation varied from 68% at 7 1C to 96% at 25 1C Manzano et al. (1999)

Light 4-Nitrophenol Light degradation (indirect photolysis) of para-nitrophenol by Rhodobactercapsulatus

Roldán et al. (1998)

4-Nitrophenol Total biodegradation of para-nitrophenol by microalgae within 5 days Lima et al. (2003)4-Nitrophenol Photocatalytic degradation of para-nitrophenol (44–80%) on nanometre size

titanium platesLi et al. (2005)

Salicylate, phenol, phenanthrene Photosynthesis-enhanced biodegradation of aromatic chemicals by algal–bacterial microcosms

Borde et al. (2003)

2,4-Dinitrophenol (2,4-D) Total biodegradation of 2,4-D with mixed culture of phototrophicmicroorganisms in the dark. Other conditions caused accumulation ofmetabolite 2-amino-4-nitrophenol (2-ANP)

Hirooka et al. (2006)

Oxygen Phenol Anaerobic degradation in digested sludge (majority monosubstitutedphenols degraded)

Boyd et al. (1983)Phenol, benzoates, phthalic acidesters, pesticides, homocyclic andheterocyclic ring chemicals,monosubstituted benzenes

Battersby and Wilson(1989)Anaerobic degradation of pyridine and quinoline.1- and 2-naphthol and

anthraquinone had inhibitory effect on digesting sludge

4-Nitrophenol Aerobic degradation of para-nitrophenol within 11 h by activated sludge Bhatti et al. (2002)

Bed form Diclofenac, bezafibrate, ibuprofen,naproxen, gemfibrozil

Flow velocity and sediment dynamics (water–sediment interactions; flatsediment vs. moving sediment) affected attenuation of acidicpharmaceuticals. Majority degraded within 2.5–18.6 days while gemfibrozildegraded after 10.5 days

Kunkel and Radke (2008)

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the complexity and dynamism/variability of environmental com-partments e.g., flowing rivers. Microbial community characterisa-tion helps to better understand the microbiology of samples thatare used as inocula in biodegradability testing and their potentialto degrade a given chemical. This information could be used tounderstand the basis of a given test result, such as a test failure, aswell as for extrapolation of laboratory test results into theenvironment from which the inoculum is collected. Optimisationof microbial inocula could also help to reduce the costs of fatestudies and risk assessments by increasing the probability ofbiodegradation and decreasing the rate of test failures, by increas-ing the reproducibility of fate studies.

Current OECD tests are conducted with variable inocula andthere is no inoculum standardisation prior to biodegradabilitytesting, which has been identified as the key factor affecting thereproducibility of chemical fate studies (Mezzanotte et al., 2005;Thouand et al., 2011). Molecular ‘omics’ tools are increasinglyaffordable, commonly used methods, which could be used along-side traditional OECD methods to study chemical biodegradationin the environment.

Application of advanced high throughput next-generationsequencing methods such as amplicon sequencing, metagenomicsand transcriptomics (Table 4) can be used for in depth character-isation of diversity and function of a community. Metagenomicsallows us to sequence genomes from complex assemblages ofmicrobes as a single unit in a culture-independent manner(Moran, 2009) and it provides an opportunity for linking microbialdiversity with function (Malik et al., 2008). Metatranscriptomics(Table 4) profiles the collective transcriptomes of a given habitat,and allows actively expressed genes (mRNA) to be retrieved andsequenced to profile active metabolic processes (Stenuit et al.,2008; Moran, 2009; de Menezes et al., 2012).

Metatranscriptomics can be applied as a supporting tool inorder to perform inoculum characterisation, to study the micro-biology and functional diversity of inocula. Such techniques could beused to establish new ways of conducting biodegradation studiesand to interpret biodegradation tests for certain chemicals, forexample by understanding the genetic potential for biodegradation

within inocula. Genetic potential for chemical biodegradation can bedetermined in an inoculum by analysing genes that are expressed bymicrobial populations during biodegradation of chemicals. Presenceof specific genes encoding enzymes responsible for breakdown ofchemicals indicates microbial potential to chemical biodegradation.This can enable comparison of chemical biodegradability studies andprediction of chemical fate. Application of microbial ecology methodsalso provides the opportunity to study chemical biodegradationpathways and to gain more insights into complex biodegradationprocesses conducted by microbial communities in a given environ-ment. Several functional gene approaches have been developed tostudy the diversity of chemical degrading microbial communities.Changes of key catabolic genes (catechol 2,3-dioxygenase, catechol1,2-dioxygenase and alkane-catabolic genes (alk)) and microbialcommunity structure were studied by Zhang et al. (2008) duringthe degradation of nonylphenol ethoxylates (NPEOs) and NP innatural water microcosms. This study demonstrates that functionalgenes may be used to monitor the potential for chemical biodegra-dation. These methods may be applied to support biodegradationtests by studying the functional diversity of inocula and its potentialfor biodegradation. Further development of such approaches couldpotentially identify the scope for adaptation within communities todegrade chemicals. Chemical fate is associated with proliferation ofdegrading bacteria which often leads to shifts (changes) in bacterialcommunity composition. Shifts occurring in microbial populationsduring biodegradation of chemical, in particular the abundance anddiversity of specific degraders can be then correlated with chemicalfate. Whereas, lack of changes in bacterial community and lowabundance of specific degraders coupled with lack of chemicalbiodegradation may suggest potential persistence of given chemical.Moreover, metagenomics, proteomics andmetabolomics allow betterunderstanding of complex changes in microbial community func-tions, detection and identification of existing and novel proteins(enzymes) and metabolites involved in the biodegradation ofchemical of interest. For example, Mackelprang et al. (2011) reportedthat metagenomics allowed examination of whole biochemicalpathways and associated processes, as opposed to individual piecesof the metabolic matrix. Iwai et al. (2010) used gene-targeted

Table 4Application and potential of high-throughput technologies in biodegradation studies.Adapted from Stenuit et al. (2008)

Technique Area of application Potential Application in fate studies and risk assessment

Fingerprintingtechniques based onamplicon sequencing

� Community structureand dynamics

Relate microbial community analysis to themetabolic function of specific groups ofbacteria

Will enable more realistic exposure assessments and enable betterextrapolation from in vitro test systems to in situ conditions

Metatranscriptomics � Community function Identification of the biodegradationpotential and the function of specificmicrobial communities

Optimisation of inoculum characterisation, which will reduce therate of test failures and increase the reproducibility of fate studiesMetaproteomics

Metabolomics

Biosensors � Presence of undefinedchemicals

� Nature andconcentration of thechemicals

� Toxicity of thechemicals for livingorganisms

� Decrease in chemicalconcentration, and/ortransformation

Monitoring ‘real-time’ the performance of abioremediation process

Understand genetic potential for biodegradation within inoculaProvide new methods and insights to study chemicalbiodegradation pathwaysProvide in silico tools to better characterise and predictbiodegradation based on chemical structure and microbialcommunities potential to degrade

Cultivation (þgenomesequencing),metagenomics

� Use of exogenousbiocatalysts

� Search for newcatabolic activities

New metabolic functions of microbialcommunities

Understanding interactions between chemicals andenvironmental factors that may modulate microbial responses

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metagenomics to study the diversity of aromatic dioxygenase genesinvolved in the turnover of more recalcitrant organic chemicals. Suchgenes can be used in the assessment of inoculum biodegradationpotential, and hence in prediction of chemical fate in the environ-ment. These approaches are useful for assessment of chemicalpersistence since they provide insight into processes such as inocu-lum adaptation and can also address evolutionary issues. Introduc-tion of new chemicals into the environment often leads todevelopment of novel chemical biodegradation pathways and isassociated with the presence of specific degrading bacteria. ‘Omics’will help us to understand such processes and their consequences forecosystems.

In metaproteomics (Table 4), complex mixtures of proteins froman environmental sample are separated and fractions of interest areanalysed by high-throughput mass spectrometry-based analyticalplatforms (Stenuit et al., 2008). Exploring the differential expressionof a wide variety of proteins and screening of the entire genome forproteins that interact with particular mineralisation regulatoryfactors is helpful to get insights into biodegradation (Vieites et al.,2009). For instance, profiles of enzymes involved in biodegradationpathways, which are expressed by microbial inocula in the presenceof a test chemical or specific metabolites, can be used to identify thepotential for chemical degradation and microbial responses tocontaminated environments. Therefore, proteomics plays an essen-tial role in determining the physiological changes of microorgan-isms under specific environmental influences, while functionalproteomics (based on the presence of enzymes involved in chemicalbiodegradation pathways) would predict metabolism of chemicalsby degrading organisms (Chauhan and Jain, 2010).

Another approach, environmental metabolomics (Table 4) hasmany advantages for studying organism–environment interactionsand for assessing organism function and health at the molecularlevel (Bundy et al., 2009). The role of metabolomics and proteo-mics in understanding mechanisms of microbial aromatic degra-dation was described by Seo et al. (2009). They reported thatmetabolomics can be used to profile degradation products of PAHsand primary metabolites in response to PAH exposures. Keumet al. (2008) studied comparative metabolic responses of Sinorhi-zobium sp. C4 during degradation of phenanthrene. Comprehen-sive profiling of metabolite profiles, including polar metabolites,fatty acids and polyhydroxyalkanoates was performed throughuntargeted metabolome analyses. Large metabolome differenceswere observed between the cultures grown with phenanthreneand natural carbon sources. Metaproteomics and metabolomicscan be used to determine the biodegradation potential of inoculaapplied in fate studies, in determination of chemical fate pathwaysin the environment, in accurate assessment of primary or ultimatebiodegradation based on metabolites. Metabolomics can beapplied to identify signature metabolites which are indicative ofchemical biodegradation (Parisi et al., 2009), and hence, can beused to predict degradation pathways and the residual concentra-tion of chemicals in receiving environments.

Biosensors also have the potential to be applied in biodegrada-tion studies (Table 4). Microorganisms are increasingly used asspecific sensing devices for measuring chemical concentrations inthe environment. These biosensors measure gene expressioninduced by the presence of a chemical of interest, which servesas a measure of its bioavailable concentration (Stiner andHalverson, 2002). As such, biosensors can detect biologicallyrelevant chemical concentrations. Such tools could be used tosupport biodegradation studies and for estimating chemical per-sistence in the environment.

Most screening and simulation OECD tests lack environmentalrealism, and this leads to difficulties with data extrapolationfrom laboratory to real world scenarios. Hence, the lack ofenvironmental realism is a key challenge for environmental risk

assessment and there is the need to develop tools that takeaccount of the complexity of exposed environments and enableassessment of site-specific effects (SCHER, SHENIHR, SCCS, 2013).Although simulation of complex environmental systems underlaboratory conditions is still challenging, better understanding ofecosystems can be achieved through the application of microbialecology and systems biology.

General knowledge about the processes operating within amicrobial community is not sufficient to determine the role ofindividual members and their interactions within the community.For biodegradation studies it is crucial to understand how microbialrelationships could be affected by the environment and newchemicals entering the environment, and vice versa (Zengler andPalsson, 2012). Also, the biodegradation processes are framed in acomplex web of metabolic and regulatory interactions which areextremely difficult to approach with traditional molecular methods.The recent accumulation of knowledge about the biochemistry andgenetics of the biodegradation process, and creation of structureddatabases, has recently opened the door to systems biology (Trigoet al., 2009). Linking microbial functions with processes occurring inthe environment can be useful for designing novel test systemssimulating environmental conditions. ‘Omics’ techniques make itpossible to understand the reasons underlying test failure andsuccess. For instance, it can be determined how laboratory condi-tions affect the inoculum potential (i.e., presence and diversity ofspecific degraders) for chemical biodegradation. ‘Omics’ could alsohelp us to understand interactions between chemicals and environ-mental factors e.g., temperature, salinity, organic matter content, andcan indicate how environmental factors might modulate microbialresponses to chemicals. This information can be applied in chemicalfate studies to simulate the right conditions for adaptation of inoculaand chemical biodegradation. Improving test conditions can help todecrease the rate of test failures and may improve laboratory to fieldextrapolations.

4.1. Advantages and disadvantages of ‘omics’

Although, high throughput techniques are emerging tools inbiodegradation studies they are still relatively new techniques.One of the biggest disadvantages of ‘omics’ is their blind applica-tion. While there is a place for discovery science using high-throughput techniques, progress is usually achieved in studies thatare applied with a clear hypothesis in mind. The major concernsarising around ‘omics’ are also linked with constant developmentof technologies and urge to bring new improvements and applica-tions by commercial enterprises. This requires constant develop-ment and education of scientists in order to be on top of thoseapproaches. Another challenge is the interpretation and manage-ment of massive amount of data, which are generated by highthroughput technologies. Therefore, development of efficient sta-tistical algorithms, pipelines and other bioinformatic tools iscrucial to obtain good quality data and to perform interpretation.However, the costs of generating high throughput data and dataanalysis are becoming more affordable which increases their avail-ability and applicability for biodegradation studies (Shendure andJi, 2008; Desai et al., 2010).

5. Conclusions

In the current review we have demonstrated the major limita-tions of present OECD biodegradation tests, and it is noticeablethat there is the need for new/modernised OECD biodegradationtest guidelines. Efforts should be made to adopt emerging techni-ques (i.e., high throughput sequencing) that would allow theapplication of novel approaches that enable inoculum

A. Kowalczyk et al. / Ecotoxicology and Environmental Safety 111 (2015) 9–22 19

characterisation and standardisation, and give insights into che-mical biodegradation pathways and improve our understanding ofchemical fate in the environment. However, the modernisation ofOECD guidelines may create several major challenges, includingthe costs of approval and application of new guidelines. It may alsoinvolve the need to continuously update regulation/guidanceregarding biodegradability tests as scientific knowledge evolves.The chemical industry and regulators may also have to facechallenges regarding the impact of novel approaches and modifiedOECD tests on the future test results, especially their interpretationand overall chemical risk assessment. The updating of currentOECD tests would also mean additional costs associated withmodernisation of research laboratories so they can adopt andimplement new test guidelines. This may result in some additionalchallenges regarding the level of expertise of researchers thatconduct these tests and the need for their further development.Introduction of new OECD guidelines would also require normal-isation of test systems and methods among all research labora-tories which perform the biodegradability testing, to assure thesame standard of quality of results, and to enable comparisonbetween biodegradation studies conducted by different laboratories.Nevertheless, there is the need for solutions that address the mainissues regarding current biodegradability testing, and application ofadvanced microbial ecology methods offers an opportunity toimprove present OECD biodegradation test guidelines.

Acknowledgements

The present work was funded by Unilever, Cefic LRi, Engineeringand Physical Sciences Research Council, AstraZeneca and Biotech-nology and Biological Sciences Research Council.

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