Contribution of microorganisms to non-extractable residue formation during biodegradation of ibuprofen in soil

Science of the Total Environment 445–446 (2013) 377–384

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Contribution of microorganisms to non-extractable residue formation during biodegradation of ibuprofen in soil Karolina M. Nowak a, c,⁎, Cristobal Girardi a, Anja Miltner a, Matthias Gehre b, Andreas Schäffer c, Matthias Kästner a a b c

UFZ, Helmholtz Centre for Environmental Research, Department of Environmental Biotechnology, Permoserstraße 15, 04318 Leipzig, Germany UFZ, Helmholtz Centre for Environmental Research, Department of Isotope Biogeochemistry, Permoserstraße 15, 04318 Leipzig, Germany Department of Environmental Biology and Chemodynamics, Institute for Environmental Research (Biology V), RWTH Aachen University, Worringerweg 1, 52074 Aachen, Germany

H I G H L I G H T S ► ► ► ►

Biogenic residue formation during microbial degradation of ibuprofen was studied. Nearly all non-extractable residues derived from ibuprofen were biogenic. Fatty acids and amino acids formed biogenic non-extractable residues and were stabilised in soil. Environmental risks of ibuprofen-derived non-extractable residues are overestimated.

a r t i c l e

i n f o

Article history: Received 31 May 2012 Received in revised form 5 December 2012 Accepted 5 December 2012 Available online 27 January 2013 Keywords: Stable isotope Ibuprofen biodegradation Non-extractable residue Fatty acid Amino acid

a b s t r a c t Non-extractable residues (NER) formed during biodegradation of organic contaminants in soil are considered to be mainly composed of parent compounds or their primary metabolites with hazardous potential. However, in the case of biodegradable organic compounds, the soil NER may also contain microbial biomass components, for example fatty acids (FA) and amino acids (AA). After cell death, these biomolecules are subsequently incorporated into non-living soil organic matter (SOM) and are stabilised ultimately forming hardly extractable residues of biogenic origin. We investigated biodegradation of 13C6-ibuprofen, in particular the metabolic incorporation of the 13C-label into FA and AA and their fate in soil over 90 days. 13C-FA and 13C-AA amounts in the living microbial biomass fraction initially increased, then decreased over time and were continuously incorporated into the non-living SOM pool. The 13C-FA in the non-living SOM remained stable from day 59 whereas the contents of 13C-AA slightly increased until the end. After 90 days, nearly all NER were biogenic as they were made up almost completely by natural biomass compounds. The presented data demonstrated that the potential environmental risks related to the ibuprofen-derived NER are overestimated. © 2012 Elsevier B.V. All rights reserved.

1. Introduction Biodegradation of organic contaminants in soils is understood as their transformation into mineralisation products, metabolites, microbial biomass and NER (Kästner et al., 1999). NER are believed to be a result of various physico-chemical interactions between a parent compound Abbreviations: 2,4-D, 2,4-dichlorophenoxyacetic acid; AA, amino acids; ASE, Accelerated Solvent Extraction; bioAA, biomass amino acids; bioNER, biogenic non-extractable residues; dw, dry weight; EA–C–irMS, Elemental analyser–combustion–isotope ratio mass spectrometry; FA, fatty acids; FAME, fatty acid methyl esters; GC–MS, gaschromatography–mass spectrometry; GC–C–irMS, gas chromatography–combustion– isotope ratio mass spectrometry; IBU, ibuprofen; MCPA, 2-methyl-4-chlorophenoxyacetic acid; NER, non-extractable residues; PLFA, phospholipid fatty acids; SOM, soil organic matter; tAA, total amino acids; tFA, total fatty acids. ⁎ Corresponding author at: UFZ, Helmholtz Centre for Environmental Research, Department of Environmental Biotechnology, Permoserstraße 15, 04318 Leipzig, Germany. Tel.: +49 341 235 1766; fax: +49 341 235 1471. E-mail address: [email protected] (K.M. Nowak). 0048-9697/$ – see front matter © 2012 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.scitotenv.2012.12.011

and/or its primary metabolites with SOM (Alexander, 2000; Dec et al., 1997; Mordaunt et al., 2005; Senesi, 1992). These interactions involved in the NER formation were demonstrated in simple humic acid-contaminant models (Bollag et al., 1992; Hatcher et al., 1993) and in abiotic soil set-ups (Alexander, 2000; Palomo and Bhandari, 2005, 2006; Pignatello and Xing, 1996). However, in soil biodegradation studies using 14C-labelled compounds, NER were analysed by quantification of 14 CO2 released from soil samples after previous extraction of contaminant residues from soil (Barriuso et al., 2008; Kästner et al., 1999). This approach is fast and accurate, but does not provide information about the chemical structure of the NER (Kästner et al., 1999). Due to a lack of this knowledge, compounds immobilised in SOM are generally considered to pose an toxicological hazard for living beings once they are released from SOM (Barraclough et al., 2005; Burauel and Führ, 2000). Recent studies on soil biodegradation with 13C6-2,4-dichlorophenoxyacetic acid (13C6-2,4-D) provided the first evidence that nearly all NER were biogenic and resulted from incorporation of 13C-label into

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microbial biomass (Nowak et al., 2011). Microbial components such as fatty acids (FA) and amino acids (AA) after cell lysis were incorporated into non-living SOM ultimately forming non-toxic biogenic non-extractable residues (bioNER). The environmental risks related to the formation of NER from 2,4-D are thus overestimated (Nowak et al., 2011); this might also be relevant for other readily biodegradable contaminants in soil. To date, it is still not known if bioNER from other biodegradable contaminants are also formed, since they were not identified and accounted for separately from the NER fraction in previous mass balance studies (Barriuso et al., 2008). Therefore, further studies on bioNER formation from other organic contaminants are needed for proper estimation of their biodegradation rates and potential environmental risks related to NER formation. We chose ibuprofen (IBU) as a model compound for our study, because it is an environmentally relevant compound (Buser et al., 1999). IBU is a one of the most commonly consumed non-prescription drug, with a global annual production of several kilotons and a high daily dose of 600–1200 mg (Zwiener et al., 2002). The main mechanism involved in the removal of IBU in sewage treatment plants is biodegradation (Jones et al., 2007; Suárez et al., 2008). The removal efficiency of this pharmaceutical was reported to be relatively high and to depend on seasonal conditions (60± 98%; Suárez et al., 2008). In spite of the high removal rate during sewage treatment, IBU was still detected in the final effluents of the wastewater treatment plants between 0.018 and 4.24 μg/L and in surface waters between 0.042 and 1.26 μg/L in various European countries (Santos et al., 2010). Studies on the acute toxicity of IBU to aquatic livings showed no lethal effects at a concentration below 10 mg/L (Cleuvers, 2003). However, chronic exposure to IBU at environmentally relevant concentrations (0.1–2 μg/L) induced reproduction damages in Medaka fishes (Han et al., 2010) and cyto-genetic effects in freshwater bivalves (Parolini et al., 2011). Although IBU is highly mobile in the aquatic environment and has a low affinity for sorption compared with other pharmaceuticals (Lin and Gan, 2011; Löffler et al., 2005; Xu et al., 2009), it accumulated in biosolids in the range of 24–548 μg/kg dw (Nieto et al., 2010; Radjenović et al., 2009). Therefore, both the irrigation of soil with reclaimed wastewater (Siemens et al., 2010) and application of biosolids to agricultural land will transport IBU into soils (Richter et al., 2007; Xu et al., 2009). IBU was reported to dissipate very quickly in agricultural soils (Lin and Gan, 2011; Richter et al., 2007; Xu et al., 2009). Recently, Richter et al. (2007) investigated IBU biodegradation in soil mass balance studies employing a radioactive tracer. In their studies, NER were formed quickly during biodegradation of 14C-labelled IBU [propionic acid-3-14C-IBU], and after 102 days reached ~50% and ~35% of 14C-IBU equivalents in clayey-silt and silty sand soil, respectively (Richter et al., 2007). NER contents reported by Richter et al. (2007) were high suggesting potential accumulation of this pharmaceutical in soils. However, to date, there is no information about their chemical composition. Based on these results from the 13C6-2,4-D biodegradation study (Nowak et al., 2011), we can assume that part of IBU residues may also be biogenic. We therefore studied the formation of 13C-labelled FA and AA during biodegradation of 13C6-IBU in soil and traced their fate over an incubation period. 2. Material and methods 2.1. Chemicals All chemicals, except where otherwise specified, were purchased from VWR (Darmstadt, Germany) at the highest quality available. Labelled 13C6-IBU (99 at.% 13C, chemical purity 98%) was purchased from Alsachim (Illkirch, France). 2.2. Soil The reference soil was taken from the agricultural long-term experiment “Statischer Düngungsversuch” located in Bad Lauchstädt,

Germany (Blair et al., 2006), where a Haplic Chernozem was cultivated continuously with crop rotation (sugar beet, summer barley, potatoes and winter wheat) and fertilised every second year with farmyard manure (30 t/ha) since 1902. The soil contained about 21% clay, 68% silt, and 11% sand. The total N was 0.17% and the organic C content 2.1% (w/w). The pH was 6.6 and the maximum water holding capacity (WHC) was 37.5%. 2.3. Soil incubation experiment with

13

C6-IBU

Three types of soil incubations were implemented: 1) unamended soil (blank, without IBU and stabilised sewage sludge), 2) soil+unlabelled IBU (control) and 3) soil amended with 13C6-IBU (biotic system). The blank and control systems provided the information on the natural abundance of 13C-FA and 13C-AA in soil. To test whether 13C-label incorporation into FA and AA is a result of biotic processes, sterile soil incubations with the 13C6-IBU (abiotic system) were also performed. Soils were sterilised using autoclave three times at 120 °C for 20 min prior to their amendment with IBU and incubation. In order to allow re-growth of microorganisms which could survive previous sterilisation steps, soils were autoclaved every third day. Soils (control, abiotic and biotic systems) were amended with stabilised sewage sludge (1.5 g dw sludge/kg dw soil) from the wastewater treatment plant “Klärwerk Rosental” located in Leipzig, Germany. IBU ( 13C-labelled or unlabelled) was dissolved in acetonitrile. To prevent killing of all microorganisms, only 10% of the total amount of soil used for each experiment was spiked and the excess of acetonitrile was air-dried under a fume hood. This spiked portion was then mixed with the rest of the soil, yielding a nominal IBU concentration of 20 mg/kg (581.5 μmol/kg) soil. However, the concentration of 20 mg/kg of soil was also confirmed by GC–MS. IBU was extracted from three soil subsamples prior to their placement into the respective bottle and incubation. Due to the low sensitivity of the 13C isotopic analytical methods it was necessary to use high concentrations both for analytical reasons and for understanding of bioNER formation in soil. The water content of the soil was adjusted to 60% of its maximum WHC and 40 g was weighed in 1000 mL Duran glass bottles. Incubation experiments were performed according to OECD guideline 307 (OECD, 2002) in the dark and at constant temperature (20 °C) for 90 days. The bottles were sampled destructively after 2, 7, 14, 28, 59 and 90 days (abiotic systems only after 28 and 90 days of incubation). The soil samples were analysed for the amount and isotopic composition of FA and AA. 2.4. FA analyses The incorporation of 13C-label derived from 13C6-IBU into two fractions of FA was analysed: the phospholipid fatty acids (PLFA) and the total FA (tFA). The PLFA represent the FA in the membranes of living cells (Zelles, 1999), whereas the tFA include storage lipids as well as the non-living, SOM-bound fraction of the FA in addition to the membrane lipids (Drenovsky et al., 2004). The PLFA fraction was extracted with a mixture of phosphate buffer, methanol and chloroform according to Bligh and Dyer (1959) and separated from neutral lipids and glycolipids by column chromatography over silica gel (Unisil, Clarkson Chromatography Products, South Williamsport, USA; Thiel et al., 2001). PLFA were derivatised using a mixture of methanol/trimethylchlorosilane (9:1; v:v) as described previously (Miltner et al., 2004). To analyse tFA, the dry soil was directly derivatised in the same way as the PLFA. After derivatisation, methylated tFA were extracted from soil with diethyl ether and purified over silica gel columns (Mallinckrodt Baker Germany, Griesheim, Germany; Miltner et al., 2004). The fatty acid methyl esters (FAME) in both fractions were identified and quantified after separation on a BPX-5 column (30 m×0.25 mm× 0.25 μm) by means of GC–MS. The isotopic composition of the individual

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FAME was determined after separation on the same column type (50 m × 0.25 mm× 5 μm) using gas chromatography–combustion– isotope ratio mass spectrometry (GC–C–irMS). The exact analytical conditions were described by Nowak et al. (2011). FA nomenclature used for the data interpretation was described in numerous papers (Kaur et al., 2005; Zelles, 1999). 2.5. AA analyses AA from protein hydrolysis were analysed for the incorporation of 13C6-IBU-derived C in two fractions: the total AA fraction (tAA) in the soil and the AA in the living microbial biomass (bioAA). tAA were determined according to Silfer et al. (1991). Soil samples were hydrolysed with 6 M HCl for 22 h at 110 °C. After hydrolysis and evaporation of the HCl, samples were purified over cation exchange resin (DOWEX 50W-X8, 50–100 mesh) as described previously by Amelung and Zhang (2001) and Nowak et al. (2011). Briefly, the samples containing the AA were washed with oxalic acid followed by 0.01 M HCl, de-ionised water and finally the AA were eluted with 2.5 M NH4OH. After purification, the carboxyl groups of AA were esterified with a mixture of isopropanol/CH3COCl and the amino groups were trifluoroacetylated with a mixture of CH2Cl2/(CF3CO)2O (Miltner et al., 2009). After derivatisation, the samples were dissolved in chloroform and the remaining impurities extracted into phosphate buffer (Ueda et al., 1989). For the determination of the AA in the living biomass (bioAA), the biomass was first extracted from the soil (Miltner et al., 2009). The cells were extracted from the soil with 1 g Amberlite IRC-748 and sodium deoxycholate/polyethylenglycol solution. Biomass pellets containing AA were further hydrolysed, purified and derivatised as described above. Two independent studies with different soils showed that the recovery of microbial biomass was in the range of 40% (Jacobsen and Rasmussen, 1992; Miltner et al., 2009). Therefore, the results show both the original data and recalculated values based on 40% extraction efficiency, and are expected to be consistent between samples although the bioAA are underestimated. However, the interpretations of bioAA were limited to the original data. The identity, quantity and isotopic composition of the AA were determined by means of GC–MS and GC–C–irMS, using the same columns as for FA analysis. The analytical conditions for AA separation using either GC–MS or GC–C–irMS was as described by Nowak et al. (2011). 2.6. Data analyses, mass balance and total bioNER content 2.6.1. 13C natural abundance A stable isotope tracer (13C) was used to investigate the formation and the fate of bioNER during biodegradation of 13C6-IBU in soil. However, the soil is also naturally abundant in 13C (~1%), therefore blank (without IBU and stabilised sewage sludge) and control (unlabelled IBU) samples were used for the correction of 13C abundance in soil with the 13C6-IBU. In addition, the 13C/12C ratios of both 13C-FA and 13C-AA in the control and blank samples remained nearly constant over 90 days of incubation. Therefore, the biological carbon isotope fractionation did not affect the presented results. The contents of 13C-FA and 13C-AA in either living biomass or in the total soil were estimated according to Lerch et al. (2009) and were shown as a percentage of the initially applied 13C6-IBU to an experiment. The isotopic enrichment of 13C-FA or 13C-AA which was altered during the derivatisation was corrected for the introduction of additional carbon as described previously by Boschker (2004) and Silfer et al. (1991), respectively. 2.6.2. Mass balance Details on the mass balances of 13C6-IBU in the soil needed for the interpretation of data on 13C-label incorporation into bioNER were presented in Girardi et al. (2013). Briefly, at the end of the

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experiment, the mineralisation amounted to 45.2% of the added label equivalents; 13.4% was detected in the solvent-extractable portion, but only 0.5% consisted of 13C6-IBU. The total NER amounted to 30%, and the recovery of 13C derived from 13C6-IBU was 88.6%. In contrast to biotic experiments, in abiotic systems, only 3.9% of 13C6-IBU equivalents was mineralised and the total NER reached 15.4% of 13C6-IBU equivalents. Solvent-extractable portion amounted to 73.6% of the initially added 13C-label and 13C6-IBU constituted 91.6% of its total content. 2.6.3. Total bioNER As mentioned above, we extracted and analysed two 13C-labelled bioNER representatives (FA and AA), which constitute part of bioNER fraction. Calculation of the total bioNER content in soil incubated with 13C6-IBU was based on the conversion factors determined in previous experiments with bacterium Cupriavidus necator JMP 134 grown on 13C6-2,4-D (Nowak et al., 2011). In these pure culture experiments, we estimated the incorporation of 13C-label from 13C6-2,4-D into FA and AA and their proportion of the total 13C in the biomass. These studies clearly showed that 13C-labelled FA represented ~5%, and 13Clabelled AA ~50% of the total 13C in the biomass (for details see also Tab. S5 in the SI in Nowak et al., 2011). 2.6.4. Statistical analyses All incubation experiments were prepared in three replicates. The results were shown as averages and the error bars represent the standard deviation of three replicates. One-way ANOVA tests were performed to identify differences in 13C-FA and 13C-AA contents in either living biomass or in the total soil. Statistical analyses were performed with the MATLAB® software. The parameters were estimated from a set of three replicate samples. Differences were regarded as statistically significant for all tests if pb 0.05. 3. Results This section focused on kinetics of 13C-FA and 13C-AA formation and their fate during biodegradation of 13C6-IBU in soil over a period of 90 days. 13C-FA and 13C-AA were extracted from the living biomass (PLFA and bioAA, respectively) and from the total soil (tFA and tAA, respectively). Total fractions of FA or AA include accordingly FA or AA in the biomass and in the non-living SOM fractions. Therefore, a difference in the total soil (tFA or tAA) and the living biomass fractions (PLFA or bioAA) was defined as the FA or AA in the non-living SOM fractions. The chemical structures and isotopic compositions of each FA and AA were inspected by GC–MS and GC–C–irMS. Owing to those analytical procedures and to an estimate of the total bioNER contents we could prove that NER formed during 13C6-IBU biodegradation were of microbial origin. Neither 13C-FA nor 13C-AA were detected in abiotic systems demonstrating their formation exclusively in biotic systems as a result of biological metabolisation. 3.1. Formation of

13

C-FA during biodegradation of

13

C6-IBU in soil

The 13C-label incorporation into PLFA was observed in the first sampling event and their contents increased significantly during the first 28 days of incubation (p b 0.05). During that time, the contents of 13C-tFA were not significantly different from 13C-PLFA (p > 0.05) suggesting that the 13C-label was assigned only to the living biomass fraction (PLFA, see Fig. 1a). When the 13C-PLFA reached the maximum on day 28 (1% of 13C6IBU equivalents) their 13C contents decreased significantly indicating the starvation of IBU degraders (pb 0.05). Thereafter, the incorporation of 13C-PLFA into non-living fraction of SOM was observed. The 13C-FA in the non-living fraction reached the maximum (0.7% of 13C6-IBU equivalents) on day 59. At the same time, the content of 13C-PLFA

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Table 1 The classification and composition of Incubation time (days)

13

13

C-label in respective PLFA during biodegradation of

C-PLFA [10

−4

; % of

13

13

C6-IBU in soil.

C6-IBU equivalents]

2

7

14

28

59

90

Saturated straight chain 14:0 15:0 16:0 18:0 20:0

n.d. 25 (±12) 74 (±138) 142 (±345) n.d.

30 (±99) 3 (±24) 378 (±655) n.d. n.d.

32 (±8) 39 (±4) 729 (±82) 135 (±48) n.d.

301 (±12) 51 (±31) 1361 (±130) 471 (±15) n.d.

82 (±95) 104 (±85) 1499 (±719) 117 (±159) 70 (±33)

5 (±1) n.d. 935 (±171) 561 (±127) 23 (±6)

Saturated methyl branched d i 14:0 i 15:0 a a 15:0 i 16:0 i 17:0 a 17:0 10 fMe 16:0 10 Me 17:0 10 Me 18:0 18:0 bbr

e n.d. n.d. n.d. n.d. n.d. 3 (±3) n.d. n.d. n.d. n.d.

n.d. n.d. 3 (±110) n.d. 20 (±126) 31 (±116) n.d. n.d. 55 (±99) 18 (±141)

7 (±1) 61 (±5) 161 (±8) 75 (±9) 42 (±19) 19.2 (±31) 120 (±12) n.d. 77 (±1) 72 (±34)

8 (±0) 387 (±64) 190 (±51) 524 (±48) 148 (±28) 110 (±30) 382 (±49) n.d. 232 (±26) 194 (±8)

22 (±13) 287 (±188) 15 (±100) 58 (±99) 99 (±97) 253 (±250) 191 (±164) n.d. 95 (±58) 62 (±46)

5 (±2) 41 (±136) 12 (±85) 49 (±57) 61 (±52) 141 (±17) 278 (±54) 16 (±4) n.d. 56 (±24)

Monounsaturated 16:1ω7cc 16:1ω5c 18:1ω9c 18:1ω7c 18:1ω5c

359 (±28) 66 (±10) 89 (±181) 68 (±243) n.d.

197 (±10) n.d. 136 (±372) 179 (±359) n.d.

856 ↑ (±42) 25 (±14) 189 (±92) 972 ↑ (±47) 58 (±0)

1041 ↑ (±99) 242 (±63) 540 ↑ (±77) 1679 ↑ (±377) 68 (±45)

232 (±162) n.d. 28 (±219) 872 (±378) n.d.

49 (±84) 39 (±57) 474 (±104) 861 (±65) 19 (±4)

Monounsaturated branched 17:1 br

n.d.

n.d.

30 (±13)

218 (±13)

n.d.

11 (±73)

Polyunsaturated 18:2 18:2ω6,9

n.d. n.d.

4 (±209) 18 (±482)

n.d. 78 (±79)

n.d. 525 (±4)

n.d. 188 (±204)

n.d. n.d.

Saturated cyclopropyl cy 17:0 cy 19:0 Total

n.d. n.d. 1498 (±509)

153 (±169) 96 (±250) 1325 (±980)

649 ↑ (±36) 89 (±46) 4516 (±312)

944 ↑ (±239) 551 (±75) 10,168 (±1417)

317 (±153) 294 (±246) 4887 (±1322)

617 (±107) 145 (±76) 4398 (±169)

a a — anteiso; bbr — branched; cc — cis isomer; di — iso; en.d. — not detectable; fMe — methyl; values in brackets (±) represent the standard deviation of the average of three replicates; values printed in bold represent characteristic values, arrows visualise increase or decrease of the respective FA compared to the preceding sampling time.

decreased to 50% of the maximum (0.49% of 13C6-IBU equivalents). Thus, on day 59, about 60% of the label in the tFA could be assigned to the non-living SOM. From day 59 onwards, the amounts of 13C-PLFA decreased slightly until the end, whereas the 13C-FA in the non-living SOM remained constant (p> 0.05). At the beginning of the incubation, the 13C-label was detected only in saturated straight-chain and monounsaturated PLFA (Fig. 1b). From days 14 to 28, the monounsaturated PLFA (in particular 16:1ω7c and 18:1ω7c; p b 0.05; see Table 1) and cyclopropyl PLFA (cy 17:0; p b 0.05) contained more 13C-label than other PLFA (except for 16:0). Within that period, a rapid increase in their 13C contents was also observed (p b 0.05). On day 14, a low incorporation of 13C into saturated methyl branched PLFA was also found and their contents then increased strongly until day 28 (p b 0.05; except for i:14:0 and a 15:0 where p > 0.05). When the 13C-PLFA reached the maximum (day 28), the 13C contents in 16:1ω7c and 18:1ω7c and in cy 17:0 were highest (p b 0.05; except for 16:0) and then they decreased significantly (p b 0.05). Furthermore, on day 28, the content of 13C in fungal polyunsaturated 18:2ω6,9 was much higher than that of on day 14 (p b 0.05). After 59 days, when the 13C-PLFA reached the plateau, the label was distributed nearly equally between monounsaturated PLFA and saturated methyl-branched PLFA. At the end, however, monounsaturated PLFA carried more 13C-label than saturated branched PLFA. Due to the fact that 13C-label was assigned only to the living biomass fraction up to day 28 (PLFA, Fig. 1a), 13C-FA were considered to be present in the non-living SOM from day 59 onwards (Fig. 1c). On day 59, saturated strain-chain FA contained most of

the 13C-label in this fraction (Fig. 1c). Monounsaturated FA contained slightly more 13C than saturated methyl-branched FA. However, at the end, 13C was distributed equally between saturated methylbranched FA and monounsaturated FA.

3.2. Formation of

13

C-AA during biodegradation of

13

C6-IBU in soil

The 13C-label was incorporated into the living biomass AA fraction (bioAA) later (day 7, see Fig. 2a) than into PLFA (day 2). The contents of 13C-bioAA increased significantly until day 28 (p b 0.05). In contrast to PLFA, the incorporation of 13C-bioAA into the non-living SOM fraction was observed much earlier (after 7 days) and continued throughout the incubation time. The contents of 13C-bioAA reached the maximum (3.25% of 13C6-IBU equivalents) on day 28 and then declined at the end (p b 0.05); a similar time course was also observed in the case of PLFA. The content of 13 C-bioAA decreased to 23% of its maximum, reaching 0.75% of 13 C6-IBU equivalents on day 90. At the end, 93% of total 13C-AA in soil was stabilised in the non-living SOM pool, finally reaching 27% of the 13C6-IBU equivalents initially added. The continuous incorporation of 13C into bioAA and tAA proves that AA contributed to NER formation in the biotic set-up, whereas no 13C-label was detected in bioAA and tAA in the abiotic controls (see Supplementary Fig. A1). The 13C abundance in the bioAA increased until day 59 and then decreased (p b 0.05), whereas in the tAA and total NER it increased progressively during 14–59 days (p b 0.05).

1.2 1.0 0.8 0.6 0.4 0.2

13 13

c

0.8 0.6 0.4 0.2 0.0 0

10

20

30

40

50

60

70

80

90

incubation time (days) Fig. 1. (a) 13C-label incorporation within tFA (●) and PLFA (▲). 13C-label distribution in lipid classes (b) PLFA and (c) FA in non-living SOM over 90 days of soil incubation with 13C6-IBU [saturated straight-chain ( ); saturated methyl-branched ( ); monounsaturated ( ); polyunsaturated ( ) and saturated cyclopropyl ( ) FA]. The bars in panels b and c show 13C-label distribution in the PLFA and in the FA in non-living fraction after the respective sampling dates: 2, 7, 14, 28, 59 and 90 days. Results were expressed in % of the applied 13C6-IBU. Error bars are the standard deviation of three replicates.

In the initial phase of 13C incorporation into the bioAA fraction (until day 7), the label was detected only in aspartate (see Fig. 2b, for details see also Supplementary Tab. A1a) and this bioAA dominated over other 13C-AA before day 28 (pb 0.05). After 14 days, glycine (0.16% of 13 C6-IBU equivalents) and lysine (0.16% of 13C6-IBU equivalents) were also relatively highly enriched in 13C. On that day, very low contents of 13C were also detected in threonine. On day 28, both 13C-aspartate (0.92% of 13C6-IBU equivalents) and 13C-glutamate (0.81% of 13C6-IBU equivalents) were the dominant bioAA, but lower 13C-label incorporation into other AA (glycine, threonine, leucine, isoleucine, proline, phenylalanine and lysine) was also observed. Interestingly, very low 13C content in the non-protein amino acid β-alanine (0.01% of 13C6-IBU equivalents) was also found on day 28; its 13C content increased rapidly by day 59 (0.05% of initial 13C6-IBU added; p b 0.05). The amounts of the dominant 13C-aspartate had decreased strongly by day 59 (pb 0.05) and on that day the label was distributed between all bioAA. At the end, 13C-label was distributed equally between the six following AA: aspartate, leucine, isoleucine, proline, phenylalanine and lysine (p>0.05). The AA were also incorporated into the non-living SOM fraction (Fig. 2c, for details see also Supplementary Tab. A1b) from day 14. Although on day 7, 13C-label was detected in aspartate, its content was negligible (≥0; p > 0.05). On day 14, the label was found only in the dominant 13C-aspartate (3.0% of 13C6-IBU equivalents, p b 0.05) and

C-non-living AA (% of initial 13 C6-IBU equivalents)

13

0.4

C-bioAA (% of initial C6-IBU equivalents)

0.6

381

35

a

30 25 20 15 10 5 0

13

0.8

0.0 1.0

13

b

1.0

C-non-living FA (% of initial 13 C6-IBU equivalents)

C-PLFA (% of initial C6-IBU equivalents)

0.0

0.2

13

a

1.4

C in AA fraction (% of initial 13 C6-IBU equivalents)

1.6

13

13

C in FA fraction (% of initial 13 C6-IBU equivalents)

K.M. Nowak et al. / Science of the Total Environment 445–446 (2013) 377–384

b 4 3 2 1 0 30

c

25 20 15 10 5 0 0

10

20

30

40

50

60

70

80

90

incubation time (days) Fig. 2. (a) 13C-label incorporation within tAA (●), bioAA (▲), and bioAA recalculated based on 40% extraction efficiency (×). 13C-label distribution (b) in the bioAA fraction and (c) in the non-living AA fraction over 90 days of soil incubation with 13C6-IBU [lysine ( ); phenylalanine ( ); glutamate (incl. glutamine) ( ); aspartate (incl. asparagine) ( ); proline ( ); isoleucine ( ); leucine ( ); valine ( ); threonine ( ); glycine ( ) and alanine ( )]. The bars in panels b and c show 13C-label distribution in the bioAA and in the non-living AA fraction after the respective sampling dates: 2, 7, 14, 28, 59 and 90 days. Results were expressed in % of the applied 13C6-IBU. Error bars are the standard deviation of three replicates.

in 13C-isoleucine (0.3% of 13C6-IBU equivalents); this is consistent with the incorporation of these AA into the biomass fraction in the initial phase of biodegradation. On day 28, the 13C-label was distributed nearly equally between all AA (p> 0.05) except for alanine, valine, β-alanine and lysine (no 13C was found). After 59 days, 13C-aspartate dominated (9.9% of 13C6-IBU equivalents; p b 0.05), but glutamate also contained a high amount of 13C (5.9% of 13C6-IBU equivalents). During that time, lower 13C-label incorporation into other AA (glycine, leucine, isoleucine, proline, phenylalanine and lysine) was also observed. At the end, all 13C-microbial-derived AA, except for threonine, were stabilised in the non-living SOM fraction to a similar extent (p> 0.05). 3.3. Total bioNER content in soil incubated with

13

C6-IBU

Calculation of the total bioNER content was based on the conversion factors for FA (factor 20) and AA (factor 2) as described previously in Section 2.6.3. From the maximum content of 1.2% (of 13C6-IBU equivalents) detected as 13C-tFA, we can thus conclude that approximately 24% of the label biomass had been incorporated into biomass intermediately. However, the recent study with 13C6-2,4-D (Nowak et al., 2011) showed that FA were continuously metabolised by soil microorganisms; their contents thus decreased over time, whereas AA remained

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Fig. 3. Conversion of the 13C-label during microbial degradation of 13C6-IBU in soil. a: General mass balance (Girardi et al., 2013) and b: new mass balance considering bioNER formation. ⁎The biogenic residues were estimated from AA using a conversion factor of 2 for proteins (for details see Sections 2.6 and 3.3).

surprisingly stable. Therefore, only AA were taken into consideration as the representative microbial biomarker for the calculation of the total content of bioNER in soil. From the final amount of 13C-tAA (27% of 13 C6-IBU equivalents), we can estimate that there were overall 54% bioNER in this experiment. 4. Discussion The results provide detailed insight into the incorporation of 13Clabel from 13C6-IBU via microbial biomass to the non-living part of SOM. Microorganisms used carbon originating from the pollutant to synthesise their biomass components, as proven by 13C incorporation into two microbial biomarkers FA and AA. After their death and cell lysis, biomass constituents such as 13C-PLFA and 13C-bioAA were distributed into the food web and finally converted into refractory non-living SOM, forming so called bioNER. At the end, bioNER represented the major fraction of NER in the 13C6-IBU biodegradation experiment. These results are in good accordance with previous studies on bioNER formation from 13C6-2,4-D (Nowak et al., 2011) and on the fate of 13C-labelled E. coli in soil (Kindler et al., 2006, 2009; Miltner et al., 2009). 4.1. Non-extractable

13

C6-IBU residues in soil are bioNER

The formation of IBU-derived NER during the first 28 days of incubation in abiotic soils was significantly lower in contrast to biotic soils (Girardi et al., 2013). At the end, their contents increased and reached 15.4% of 13C6-IBU equivalents but only 3.9% of 13C6-IBU equivalents

was transformed into 13CO2. However, no 13C-label incorporation into FA and AA was observed in the abiotic soil systems. The formation of NER in abiotic soils thus was time-dependant and could have occurred via “slow” sorption to or entrapment by SOM of parent molecules or their primary metabolites (Alexander, 2000). Soil used in these abiotic experiments was autoclaved three times. This procedure might have changed its structure, physicochemical properties and thus also the sorption potential of 13C6-IBU (Berns et al., 2008; Shaw et al., 1999). However, the IBU present in the soil can be either sorbed or biodegraded. Biodegradation in the biotic soils will therefore result in lower sorption of the compound. Thus, if NER formation by sorption is estimated based on the comparison with an abiotic soil, it will be overestimated if the compound is significantly biodegraded. The incorporation of the 13C-labelled microbial components (FA and AA) into the non-living fraction of SOM and the high extent of IBU mineralisation (45.2% of 13C6-IBU equivalents; see Fig. 3a) thus indicated the relevance of microorganisms for NER formation in biotic soils. At the end, 27% of 13C6-IBU equivalents were found in tAA, corresponding to a final amount of 54% for bioNER (for details see Section 3.3). However, the total 13C-NER content determined by Girardi et al. (2013) was much lower (30% of 13C6-IBU equivalents, see also Fig. 3a) than this estimated amount of bioNER. This could be a result of partial extraction of 13C6-IBU-derived bioNER prior to NER determination by Elemental Analyser–Combustion–isotope ratio Mass Spectrometry. At the end, 13.4% of 13C6-IBU equivalents were detected in the solvent-extractable portion, but 13C6-IBU constituted only 0.5% of its total content (Girardi et al., 2013). The remaining parent 13 C-IBU and its metabolites were extracted from soil using accelerated solvent extraction method, which is considered to be a ‘harsh’ extraction method (Northcott and Jones, 2000). The chemical composition of other compounds in this solvent-extractable fraction was not analysed. Therefore, we can assume that a certain part of bioNER might have been extracted, in particular living biomass compounds prior to their stabilisation in the SOM pool. In addition, the conversion factor to calculate the bioNER from the protein content is just an estimate based on a 50% protein content in microbial biomass. An actual protein content of the biomass slightly higher than this 50% would result in an overestimate of the amount of bioNER. The continuous decrease of the content of extractable IBU residues of unknown composition (Fig. 3a) and simultaneous incorporation of living biomass components into non-living SOM indicated a progressive stabilisation of bioNER. Overall, at the end, the remaining unmineralised IBU-derived NER could be assigned to its biogenic counterpart and contained only microbial cell constituents (see Fig. 3a and b). These results support our earlier study on NER formation from 13C6-2,4-D in soil, where biomass compounds were finally converted into bioNER after 32 days (Nowak et al., 2011). However, the formation of bioNER from 13C6-2,4-D was more rapid (Nowak et al., 2011) in comparison to that of 13C6-IBU. This may be explained by the higher solubility in H2O of 2,4-D (600 mg/L; Villaverde et al., 2008) than IBU (21 mg/L; Yalkowsky and Dannenfelser, 1992). Therefore, 2,4-D is present in the water-soluble fraction of soil at higher concentration than IBU, resulting in a better availability of 2,4-D than IBU for degrading microbes (Semple et al., 2004) and thus faster biometabolisation. Furthermore, 2,4-D is degraded easily by a wide range of microorganisms, even in pristine soils (Fulthorpe et al., 1996). In addition, 2,4-D degrading microbes could have been present in soil used for our experiments, since it had been treated previously with pesticides structurally related to 2,4-D (dichlorprop and MCPA, Merbach, UFZ personal communication, 2010). The IBU-derived NER content of 50% (propionic acid-3- 14C-IBU equivalents) in clayey-silt soil (Richter et al., 2007) was in the same range as the amount of bioNER found in the present work (54% of 13C6-IBU equivalents). Their formation, however, was more rapid (maximum after 4 days) in comparison to the present study (maximum after 59 days). This difference could be related to different labelling

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positions of the IBU tracer compounds used in the two studies ( 14C in the propionic acid side chain (Richter et al., 2007) vs. in the ring in our study). Murdoch and Hay (2005) demonstrated that all C atoms in the propionic side chain of IBU were lost early during degradation by bacteria isolated from a waste water treatment plant. 4.2. FA and AA as biogenic components of 13

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C6-IBU-derived NER in soil

C-label was incorporated first into PLFA then into bioAA suggesting the preference of 13C-PLFA synthesis over 13C-bioAA in the initial biodegradation phases in soil. This is in good accordance to previous study (Nowak et al., 2011). The observed increase or decrease in 13C contents of PLFA and 13C-bioAA in time course was dependant on the availability of substrate for microorganisms. The contents of 13C in both biomarkers increased rapidly until day 28, when the 13C6-IBU was still detectable (Girardi et al., 2013; see also Fig. 3a), and thus served as a potential carbon source for growing microorganisms. When available 13C6-IBU exhausted, the 13C-FA and 13 C-AA in the living biomass declined and were incorporated continuously into non-living SOM, ultimately forming bioNER. As outlined in the Results section, the 13C-label derived from 13C6-IBU was incorporated into diverse classes of lipids (monounsaturated, saturated methyl-branched and cyclopropyl PLFA); its abundance within the lipid classes was dependant on the time course. Monounsaturated PLFA were the dominant 13C-PLFA until day 28, whereas saturated methyl-branched PLFA were also highly enriched in 13C in later phases of the incubation period. The first class of PLFA is characteristic for Gram-negative bacteria, the second one for Gram-positive bacteria (Kaur et al., 2005). Similar results were observed in the 13C6-2,4-D biodegradation, where Gram-negative bacteria were also initial degraders (Nowak et al., 2011) Gram-positive bacteria were also found later in 13C6-2,4-D study demonstrating their relevance in indirect bioNER formation from 13C6-2,4-D via the refixation of 13CO2 released from the labelled herbicide (Nowak et al., 2011). In addition to bacteria, fungi also participated in the turnover of 13C derived from IBU as shown by the 13C found in the fungal biomarker PLFA 18:2ω6,9. The continuous decline of 13C-PLFA in the living biomass and the increase of 13C incorporation into cyclopropyl PLFA throughout the incubation time clearly indicated starvation of the bacteria (Kaur et al., 2005). With respect to incorporation of 13C-label into AA, aspartate was the first and the dominant 13C-AA found in the initial phase of 13C6-IBU degradation (up to day 28). The fast appearance of the label in aspartate suggested that IBU-derived 13C was incorporated into bioAA via heterotrophic CO2 fixation (Miltner et al., 2004). Aspartate is a precursor for lysine and isoleucine (Feisthauer et al., 2008), therefore their lower 13C amounts were also found in the initial phase of incubation. This is in contrast to the study with 13C6-2,4-D, where 13C was initially incorporated into glutamate (Nowak et al., 2011). Aspartate might also have been formed in the 2,4-D study in a very early biodegradation phase of this contaminant, but this phase might have been overlooked because 2,4-D was degraded fast, and the first sampling was only this phase. When the 13C-bioAA started to decline and 13C6-IBU was nearly depleted, 13C-glutamate was also formed in soil incubated with 13C6IBU. The high contents of 13C-glutamate found would point to a different degradation pathway for IBU-derived 13C in the later phases of the experiment. Interestingly, the non-protein AA 13C-β-alanine which was found on days 28 and 59, is a decomposition product of aspartate. The ratio of aspartate to β-alanine indicates the intensity of SOM decay (Dauwe and Middelburg, 1998). On day 28 this ratio was ~100, but it decreased to 5.4 on day 59, again indicating starvation of microorganisms, similar to the cyclopropyl PLFA. A lower label incorporation into other AA (proline, alanine, glycine, threonine, leucine, phenylalanine and valine) was also observed, as in the 13C6-2,4-D study. The decrease in the 13C-PLFA and 13C-bioAA contents observed at the end of 13C6-IBU study is in good accordance to previous biodegradation experiment with 13C6-2,4-D (Nowak et al., 2011). However, 13C-FA

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in the non-living SOM fraction remained constant until the end which is in contrast to 13C6-2,4-D study, where their contents decreased. This could be assigned to the slower turnover kinetics of 13C6-IBU (estimated degradation half-life 130 days; Girardi et al., 2013) than of 13C6-2,4-D (degradation half-life 38 days; Girardi et al., 2013). Slower biodegradation of 13C6-IBU than that of 13C6-2,4-D resulted in later incorporation of 13C into PLFA and their continuous transfer to the non-living SOM fraction until the end of the experiment. However, the content of 13C-FA in this fraction can be assumed to decrease over time in a longer incubation experiment. 5. Conclusions The results showed that microbial biomass is a source for NER formation during biodegradation of 13C6-IBU in soil by incorporating cellular components (e.g. FA and AA) carrying the isotope label into SOM. These results are supported by data from a similar experiment with the pesticide 13C6-2,4-D in the same soil (Nowak et al., 2011). In both cases, the amount of NER can be almost completely explained by bioNER that are derived from the decay of the intermediately formed microbial biomass. This contradicts the generally accepted view that NER mainly consist of the parent compounds or their xenobiotic metabolites (Barriuso et al., 2008). However, bioNER can only be formed when the organic contaminant is actually biodegraded with incorporation of carbon into microbial biomass. BioNER in soil resemble natural products and are thus explicitly excluded from the IUPAC definition of NER (Roberts, 1984). Therefore, in future mass balances of a contaminant in soil it is necessary to consider a possible formation of bioNER and to distinguish between xenobiotic derivedNER from its biogenic counterpart. Together with a previous study (Nowak et al., 2011), this study clearly showed that formation of bioNER is relevant for the readily biodegradable compounds 2,4-D and IBU. Nevertheless, studies on bioNER formation still need to be complemented by studying other group of chemicals, in particular those which display slower degradation kinetics. Acknowledgements The authors thank the European Commission for funding the RAISEBIO Project (Contract: MEST-CT-2005-020984) under the Human Resources and Mobility Activity within the 6th Framework Programme, in particular the fellowships of K. M. N and C. G. We also thank O. Carranza-Diaz (UFZ, Department of Analytics) for his help in the statistical analyses and U. Günther and F. Bratfisch (UFZ, Department of Isotope Biogeochemistry) for their valuable assistance in the compound-specific isotope analysis. In addition, three anonymous reviewers are also gratefully acknowledged for their scientific advices. Appendix A. Supplementary data Supplementary data to this article can be found online at http:// dx.doi.org/10.1016/j.scitotenv.2012.12.011. References Alexander M. Aging, bioavailability, and overestimation of risk from environmental pollutants. Environ Sci Technol 2000;34:4259–65. Amelung W, Zhang X. Determination of amino acids enantiomers in soil. Soil Biol Biochem 2001;33:553–62. Barraclough D, Kearney T, Croxford A. Bound residues: environmental solution or future problem? Environ Pollut 2005;133:85–90. Barriuso E, Benoit P, Dubus IG. Formation of pesticide nonextractable (bound) residues in soil: magnitude, controlling factors and reversibility. Environ Sci Technol 2008;42:1845–54. Berns AE, Philipp H, Narres HD, Burauel P, Vereecken H, Tappe W. Effect of gammasterilization and autoclaving on soil organic matter structure as studied by solid state NMR, UV and fluorescence spectroscopy. Eur J Soil Sci 2008;59:540–50.

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