ARTICLE pubs.acs.org/est
Cyclic Volatile Methylsiloxane Bioaccumulation in Flounder and Ragworm in the Humber Estuary Amelie Kierkegaard,*,† Roger van Egmond,‡ and Michael S. McLachlan† † ‡
Department of Applied Environmental Science (ITM), Stockholm University, SE-106 91 Stockholm, Sweden Safety and Environmental Assurance Centre, Unilever Colworth Laboratory, Sharnbrook, Bedfordshire MK44 1LQ, U.K.
bS Supporting Information ABSTRACT: Cyclic volatile methylsiloxanes are being subjected to regulatory scrutiny as possible PBT chemicals. The investigation of bioaccumulation has yielded apparently contradictory results, with high laboratory fish bioconcentration factors on the one hand and low field trophic magnification factors on the other. In this study, octamethylcyclotetrasiloxane (D4), decamethylcyclopentasiloxane (D5), and dodecamethylcyclohexasiloxane (D6) were studied along with polychlorinated biphenyls (PCBs) in sediments, ragworm, and flounder from six sites in the Humber Estuary. Bioaccumulation was evaluated using multimedia bioaccumulation factors (mmBAFs) which quantified the fraction of the contaminant present in the aquatic environment that is transferred to the biota. PCB 180, a known strongly bioaccumulative chemical, was used as a benchmark. The mean mmBAF of D5 was about twice that of PCB 180 in both polycheates and flounder, while for D4 it was 6 and 14 times higher, respectively. The mmBAF of D6 was a factor 510 lower than that of PCB180. The comparatively strong multimedia bioaccumulation of D4 and D5, even in the absence of biomagnification, was explained by both compounds having a >100 times stronger tendency to partition into lipid rather than into organic carbon, while PCB 180 partitions to a similar extent into both matrices.
’ INTRODUCTION Cyclic volatile methylsiloxanes (cVMS) are odorless liquids used as precursors in the production of silicone polymers and as solvents or fragrance carriers in personal care products and cleaning agents. Octamethylcyclotetrasiloxane (D4), decamethylcyclopentasiloxane (D5), and dodecamethylcyclohexasiloxane (D6) are three cVMS with large emissions to the environment; they are estimated to be 1.5, 17, and 2.2 kilotonnes per year, respectively, in Europe alone.13 The vast majority of the emissions are to air, but a small fraction is released to wastewater, whereby some of the cVMS reach recipient waters. There is concern that cVMS may be persistent in the recipients and that they bioaccumulate in aquatic food webs. This has led to regulatory scrutiny of these chemicals in several jurisdictions including Canada, where it has been classified as a substance that requires further assessment, and the European Union, where they are under consideration for regulation as possible vPvB (very persistent, very bioaccumulative) substances.13 Assessing the bioaccumulation of cVMS has proven to be complex. The high log KOW values of the chemicals (6.49, 8.03, and 9.06 for D4, D5, and D6, respectively13) indicate that they are hydrophobic and have the potential to bioaccumulate in aquatic organisms. Laboratory measurements are difficult because the high Henry’s law constants (1.2, 3.3, and 4.9 106 Pa m3 mol1 at 25 °C for D4, D5, and D6, respectively13) r 2011 American Chemical Society
make it difficult to control exposure to test organisms via water. For D4 and D5 the best available bioconcentration factors (BCFs) for fish (12 400 and 7060, respectively1,2) are above regulatory thresholds for bioaccumulation. This has motivated the collection of field data, and cVMS have been detected in biota from the vicinity of wastewater discharges46). The determination of bioaccumulation factors has not been possible though due to the lack of sampling and analytical methods able to quantify cVMS in water. A recent study instead focused on studying biomagnification; in a food web consisting of 2 benthic invertebrates and 15 fish species, trophic magnification factors (TMFs) 1 is the best criterion for establishing whether a chemical is a bioaccumulative hazard, none of the cVMS are bioaccumulative.7 This stands in contradiction to the conclusion from the fish bioconcentration studies. Thus, more work is required to establish the bioaccumulative properties of cVMS. To address this need, we returned to a more fundamental exposure hazard based definition of bioaccumulation, namely Received: March 7, 2011 Accepted: June 9, 2011 Revised: June 7, 2011 Published: June 09, 2011 5936
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that bioaccumulation is the fraction of the chemical present in the environment that has accumulated in the organism. From this perspective, bioaccumulation is assessed by taking the quotient of the amount of chemical in an individual organism (mCorg) and the amount of chemical in the environment (mCenv).8 This is the basis of the multimedia bioaccumulation factor (mmBAF), which in addition normalizes mCenv to the area or spatial extent of the environment (Aenv).9 To study the aquatic bioaccumulation of cVMS, we choose to restrict the environment to the aquatic compartments, namely the water column and the underlying surface sediment. The atmosphere’s influence on the fate of cVMS in the aquatic environment is primarily as a sink,10 and hence it can be neglected in assessing the aquatic bioaccumulation. This gives mmBAFaquatic ¼ ¼
mCorg Aenv mCenv mCorg Aenv m2 organism-1 mCwater + mCsed
ð1Þ
To employ the mmBAF concept for assessing bioaccumulation, one needs to calibrate it. A basis must be established for deciding whether an mmBAF value indicates that the chemical is highly bioaccumulative or not. Since there are few published values of mmBAF on which to build such a calibration, a benchmarking approach is preferable. Here the mmBAF of the chemical of interest is compared with the mmBAF of a chemical that is known to be strongly bioaccumulative. If the mmBAF of the chemical of interest and the benchmarking chemical have a similar magnitude, then the chemical of interest is also bioaccumulative; if the mmBAF of the chemical of interest is much lower, then it is not bioaccumulative. Since the concentrations of the chemical of interest and the benchmark chemical are determined in the same samples, this approach also greatly reduces a number of sources of variability and uncertainty in the mmBAF calculation such as the size of the organism, individual and site specific differences in bioaccumulation behavior, and the representativeness of the samples collected. Furthermore, employing a comparative approach to assessing bioaccumulation allows one to partially or completely circumvent many of the problems encountered with the customary approach based on absolute values of bioaccumulation metrics. It is no longer necessary to interpret the potential impact of, e.g., temperature, sorption to environmental media, organism composition, or food web structure on the measured value of the bioaccumulation metric; the bottom line is whether bioaccumulation of the test chemical is greater or less than bioaccumulation of the benchmark. In this study these concepts were applied to study the bioaccumulation of cVMS in a food chain in the Humber Estuary. Samples of sediment, common ragworm (Hediste diversicolor), and flounder (Pleuronectes flesus) were collected from six different locations and analyzed for cVMS as well as for polychlorinated biphenyls. The bioaccumulation of cVMS was evaluated according to the multimedia definition of bioaccumulation and employing PCB congeners as benchmark chemicals.
’ METHODS Study Area. Located on the east coast of England, the Humber Estuary drains approximately 20% of England’s surface area, and its watershed is home to more than 11 million people (see Figure
S1, Supporting Information). Price et al.11 estimated a wastewater treatment plant effluent flux of 2.4 mg cap1 d1 of D5 from the use home and personal care products containing D5. Assuming this is a reasonable estimate, the total Humber catchment would receive a maximum annual load of D5 of around 10 tonnes. The load reaching the estuary would be less due to adsorptive losses to riverine sediment and volatilization. Within the Humber Estuary, fish appear to have a general diet (dominated by mysids and gammarids), thus producing a complex food web. Flounder is one of the major fish species found in the estuary. It uses intertidal areas to feed and is seasonal in its distribution, with the highest abundance in June and July and no recorded occurrence in the winter.12 Hediste diversicolor (common ragworm) is the polycheate that could be obtained in the greatest numbers, at a range of salinities, within the Humber Estuary. It feeds as an active predator and omnivore and behaves both as a deposit-feeder as well as a secondary filter-feeder, collecting food near the opening of its burrow or via trapped particles on a secreted mucous filter.13 It therefore ingests comparatively large amounts of sediment relative to its body weight. Sampling. Samples of flounder, ragworm, and sediment were collected at six intertidal sites in the lower estuary between 24 September and 15 October 2009 (Figure S1) by the Institute of Estuarine & Coastal Studies at the University of Hull. Details of the sampling sites and times are given in Table S1, Supporting Information. Flounder were sampled with nets deployed during one tidal cycle. Most fish were still alive when the nets were emptied. When available, flounder of 170190 mm were selected; otherwise, other sizes were taken (no flounder were obtained at Skeffling). They were rinsed in estuary water to remove excess sediment, wrapped in aluminum foil, sealed in polyethylene bags, transferred to the laboratory in a cool box, and frozen at 17 °C. In Stockholm, the fish were partly thawed, and two dorsal skinfree fillets were excised in a clean-air fume hood, packaged in aluminum foil, and refrozen in vacuum-sealed PE pouches for later analysis (one fillet for cVMS and one for PCBs). Fish homogenates were not prepared due to the risk of sample contamination. Ragworms were sampled from undisturbed sediment at low tide. Three locations at each sampling site were chosen, and 50 individuals were sampled at each location using forceps. They were placed in plastic buckets containing estuary water and a thin layer of site sediment for transport to the laboratory cold room (10 °C). There, each individual was rinsed in site water, transferred to a 100 mL glass beaker containing 30 mL of sampling site water, and allowed to depurate for 24 h. Thereafter, they were blotted on filter paper, and whole organisms were wrapped in packages of 510 individuals in aluminum foil, sealed in polyethylene bags, and frozen at 17 °C. For each sampling location and site, some packages were analyzed for cVMS and others for PCBs. Sediments were collected within 1 m of the 3 polychaete collection locations at each site. A clean stainless steel spoon was used to remove the top layer (12 cm depth) of undisturbed sediment, which was transferred to a 500 mL glass jar. The jars were stored at 10 °C in the dark. Due to the high volatility of cVMS, there is a high risk of sample contamination via the air. This was checked for using field blanks consisting of pouches of polyester fabric (∼2 2 cm) filled with ∼60 mg ENV+ sorbent (a hydroxylated 5937
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Environmental Science & Technology polystyrene-divinylbenzene copolymer purchased from Biotage AB, Uppsala, Sweden). The pouches were rinsed repeatedly in hexane and dried with nitrogen (precleaned by passage through ENV+) before use. During sediment sampling, a glass vial containing a pouch was opened when sampling began, held close to the sample during the sampling procedure, and then resealed. During ragworm depuration, a pouch was placed in a beaker containing pure water, allowed to stand for 24 h together with the other depuration beakers, transferred back to a glass vial, and frozen. During ragworm blotting a glass vial containing a pouch was placed on the working surface, opened when blotting began, and closed when blotting of all of the ragworms was finished. Analysis of cVMS. Extraction of the biological samples was performed with a purge and trap method recently described in ref 14 with an extraction time of 24 h. The sediment samples were analyzed both as raw extracts and after an additional cleanup using ENV+ sorbent. About 10 g of dewatered sediment was extracted in a 50 mL glass tube with 10 mL of acetone (Lichrosolve,Merck) and 2 mL of pentane (Chromasolv, Sigma-Aldrich) containing the surrogate standards. The tube was rotated for 30 min and centrifuged and the liquid phase transferred to a new tube. The extraction was repeated, and the liquid phases were combined. After the addition of 20 mL of Milli-Q water, the pentane extract was centrifuged and an aliquot taken for GC/MS analysis. These results are referred to as the raw extract. The remaining pentane extract was transferred to a glass flask and purged with purified nitrogen at room temperature. The gas stream leaving the vial was passed through a sample trap with 20 mg of ENV+. After 20 h the ENV+ trap was removed and eluted in the same way as the biota samples above. Duplicate sediment samples were analyzed before and after ENV cleanup. The field blank pouches were dried with a stream of ENVfiltered nitrogen and extracted with 1 mL of n-hexane containing the surrogate standards. An aliquot was taken for GC/MS analysis. Instrumental analysis was performed on a Trace GC Ultra (Thermo Electron Corp.) equipped with a MD800 MS detector (Fisons Instruments SpA) using electron ionization (EI). A 5 μL portion of the extract was injected in a large volume splitless injector at a temperature of 220 °C. The GC temperature program and ions monitored are supplied in ref 14. A procedural blank and a control sample were analyzed with every extraction round of 8 samples. The control sample was a herring homogenate stored frozen in portions of 10 g. The extraction and all handling of the ENV+ traps were performed in a clean air cabinet. Details of other measures taken to reduce contamination during sample preparation and instrumental analysis are described elsewhere.14,15 The LOQ was calculated from the mean of the procedural blanks run with the samples (n = 4, 6, and 10 for sediment, ragworms, and flounder, respectively) plus 10 times their standard deviation. Due to the limited number of blanks used in the sediment analysis an additional criterion of 5 times the cVMS content in the procedural blank was used for the sediment LOQ. The results were not blank corrected. Analysis of PCBs. The analysis was based on the German ASU method 00.00.38.16 Briefly, the sample was mixed with sea sand/ Na2SO4, 13C-labeled standards of the analytes were added, the sample was extracted in n-hexane/dichloromethane (1:1) using accelerated solvent extraction, ASE, the extract was cleaned up on a mixed silica gel column followed by an ALOX column, and the purified extract was analyzed using GC/HRMS. The
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concentrations of PCBs 138, 153, and 180 were determined. The LOQ was defined as three times the method blank. The analyses were performed by the Oekometric GmbH.
’ RESULTS AND DISCUSSION Results of the Quality Assurance. All three cVMS were detected in each of the sediment samples. D5 and D6 were above the LOQ in all samples while D4 was below the LOQ in all samples. The amount of D5 in the samples was >900 times above the levels in the sediment field blank (Table S2, Supporting Information). The median difference between the duplicate measurements made for each sediment sample was 2% and 1% for D5 and D6, respectively, analyzing the raw extract (see Table S2). This indicated that the D5 and D6 data for sediment were of high precision. Using the raw extract is a simple and fast method, while the ENV+ cleanup provides cleaner extracts and better protection of the instrumental system. The median difference between the mean concentration measured after cleanup with ENV+ and the raw extract as such was 2% for D5 and 3% for D6. The good agreement for D6 is noteworthy considering the low recovery of less than 10% after ENV+ cleanup, and provides an indication of the effectiveness of the surrogate standard procedure used to correct for recovery. The data from the raw extract method were used in this study. D5 and D6 were above the LOQ in all of the ragworm samples, while D4 was below the LOQ in 11 of the 19 samples analyzed (see Table S3, Supporting Information). In flounder, D5 was above the LOQ in all samples, while for D4 and D6 just 9 and 2 of the 34 samples were above the LOQ (see Table S4, Supporting Information). The median difference in wet weight normalized concentrations between the replicate ragworm samples (n = 3) was 38% for D5 and 13% for D6. The corresponding median for the replicate flounder samples (n = 7) was 49% for D5. This indicates that the precision of the biota data was lower than the sediment data. Some of this is likely due to variability, for instance in lipid content, between worms from a given site or between muscle tissue samples from the same flounder, as the method repeatability was considerably lower (relative standard deviation of 12% (D5) and 15% (D6) in a control sample of herring homogenate).14 More information on the quality assurance of the data can be found in ref 14. The D4 contents of the depuration and blotting field blanks were similar to the D4 contents of unexposed blank pouches (see Table S3), and hence provided no evidence of significant contamination of the ragworms during sampling. The same was true for D5 and D6 in the depuration field blanks, but the blotting field blanks showed somewhat elevated levels. The D5 content of the ragworm samples was >80 times higher than the content of the blotting field blank. However, for D6 the difference was only a factor 1.522. It was decided not to censor any of the D6 data as the blank was exposed to the air ∼30 times longer than the individual worms, but there is a possibility of some positive bias in the samples with low D6 concentrations. The PCB concentrations in the sediments were all at least a factor of 10 above the method LOQ. The PCB levels in the biota were low, and only 14 flounder were analyzed as the remaining fish were too small to allow quantification. For the ragworms, 96% of the data points were above the method LOQ, while for the flounder only 54% were, whereby the frequency varied between congeners, being greatest for PCB 180 (79%). 5938
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Figure 1. Concentration in ragworms versus concentration in sediment for D5 and D6. Data from 6 sites in the Humber Estuary.
Levels of cVMS. The cVMS and PCB concentrations are given in Tables S2S4. The D5 concentrations in sediment ranged between 60 and 260 ng g1 dry weight (d.w.) (26008700 ng g1 OC) while the D6 concentrations ranged between 30 and 95 ng g1 d.w. (13003200 ng g1 OC, Table S2). These concentrations are higher than have been reported for the Rhine,1,2 the Arctic,6,17 and a range of lakes and coastal areas in the Nordic countries.18,19 They are comparable to the levels reported in Cardiff Bay2 the inner Oslofjord,5 Stockholm Harbour, Helsinki Harbour, Roskilde,19 Himmerfj€arden,18 and Lake Pepin,4 all areas that, like the Humber Estuary, are highly impacted by STP effluents and, in the case of Cardiff Bay, discharges from cVMS production facilities. A clear spatial gradient along the estuary was observed, with the highest concentrations at the most upstream station and the lowest concentrations at the most seaward station. The concentration gradient through the estuary is likely a reflection of the dilution of the river water as it moves down the estuary. The concentrations of D4 were less than 12 ng g1 d.w. This is consistent with the literature, which contains only sporadic reports of D4 detection in sediment. The D5 concentrations in the ragworms ranged between 51 and 760 ng g1 wet weight (ww), while the concentration in flounder muscle ranged between 12 and 300 ng g1 ww. The D6 concentrations in ragworms ranged between 2.5 and 27 ng g1 ww. D4 and D6 were quantified at low ng g1 ww levels in several flounder samples, and D4 was quantified in 8 of 19 polycheate samples at levels up to 20 ng g1 ww. These concentrations are higher than have been reported in fish from the Baltic and northeast Atlantic,13 from Sweden,18 and from other Nordic countries (with the exception of cod liver from the inner Oslofjord, but including flounder filet from that site).5 They are comparable to the levels reported in a range of fish species from Lake Pepin and in several fish species from the Rhine River close to the Dutch border.13 The variability in the D5 concentrations in the ragworms from a given site was a factor 2.7 at the most upstream site (Chowder Ness) and 1.251.60 at the other locations. The variability for flounder was even greater, amounting to a factor >2.5 at 4 of 5 stations. In most cases a single flounder sample with a high concentration contributed to the high magnitude of the range. As the flounder is not a stationary fish, this variability is likely due to differences in contamination of the environments that they had been exposed to as well as from contributions by factors such as age and feeding habits. There was a clear correlation between the concentrations in ragworms and in sediment for both D5 and D6 (Figure 1). This
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suggests that there was a direct relationship between the cVMS levels in the organisms and in the sediment. As to be expected for a nonstationary fish, no correlation between the D5 concentrations in flounder and sediment was found. The PCB concentrations in the sediments were relatively uniform throughout the estuary; e.g., for PCB 180 the highest concentration was 33 ng g1 OC at the most upstream station and ranged between 21 and 25 ng g1 OC at the other 5 stations. The within-site variation in the concentrations in ragworms was greater for the PCBs than for D5. There was no correlation between the PCB 180 concentration in ragworms and the PCB 180 concentrations in sediment, which may be partly attributable to the small range in concentration in sediment as well as the high within-site variability in the ragworms. The within-site range in the concentrations in the flounder was less for PCB 180 than for D5, but there were considerably fewer data points. Bioaccumulation of cVMS. As outlined above, cVMS bioaccumulation was assessed using the multimedia bioaccumulation factor. With eq 1, the mass of chemical in the organism (mCorg) or environmental compartment (mCwater or mCsed) can be replaced by the concentration of the chemical (CCorg, CCwater, or CCsed) multiplied by the organism’s or compartments mass (morg, mwater, or msed): mmBAFaquatic ¼
morg CCorg Aenv m2 organism-1 ð2Þ mwater CCwater + msed CCsed
Since the concentration in water had not been measured, it was estimated from the sediment/water partition coefficient, which in turn was estimated from the product of the organic carbon/ water partition coefficient KOC (L kg1) and the organic carbon content of the sediment fOC (kg kg1). The density of water Fwater was used to convert the concentration in water from volume to mass normalized units. morg CCorg Aenv
mmBAFaquatic ¼
msed CCsed 1 +
mwater =msed Fwater fOC KOC
! m2 organism-1
ð3Þ The mmBAF of the cVMS was normalized to the mmBAF of a benchmarking chemical of known bioaccumulation behavior. The mmBAF benchmarked ratio (Bratio) can then be written as Bratio ¼
mmBAFchemical mmBAFbenchmark
ðmorg CCorg Aenv Þ ¼ ðmorg CBorg Aenv Þ
CCorg ¼ CBorg
CBsed CCsed
msed CBsed
mwater =msed 1+ Fwater fOC KOCB
!
! mwater =msed msed CCsed 1 + Fwater fOC KOCC ! mwater =msed 1+ Fwater fOC KOCB ! ð4Þ mwater =msed 1+ Fwater fOC KOCC
where CBorg and CBsed are the concentrations of the benchmarking chemical in the organism and the sediment and KOC-C and 5939
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Figure 2. Logarithm of the Bratio (see text for definition) of D4, D5, and D6 in ragworms and flounder. The Bratio was calculated from cVMS and PCB 180 concentrations measured in ragworms from the same site and location or in the same flounder. A log Bratio of 0 means the chemical transfer to the organism from the aquatic environment is the same as for PCB 180. The mean and standard error of the log Bratio are shown.
KOC-B are the KOC values of the chemical of interest and the benchmarking chemical, respectively. PCB 180 was chosen as the benchmark chemical for the following reasons: (a) it is known to be a highly bioaccumulative PCB congener, and BSAF values are known to be high compared to other organic contaminants;20 (b) of the PCB congeners measured it has the highest sediment/water partition coefficient, and hence the largest proportion associated with sediment in the aquatic environment; (c) it had the highest detection frequency in the flounder (79%). Bratio was calculated using CCorg and CBorg from the same organism (fish) or sample location (ragworms) and CCsed and CBsed from the same sediment sample, whereby the biota sample and the sediment sample were from the same site. mwater/msed was assumed to be 50 (corresponding to a water depth of 5 m and a sediment bed depth of 0.1 m with a dry matter content of 1 kg L1). KOC was set to 1.7 104 L kg1 for D4,1 1.5 105 L kg1 for D5,2 6.5 105 L kg1 for D6,3 and 3.98 106 for PCB 180.21 The Bratio values were insensitive to the value of mwater/ msed; increasing it by an order of magnitude decreased the mean Bratio (see below) for D4 by a factor 1.9 and had a negligible effect for D5 and D6. This is because the sediment contained the majority of the chemical inventory in the aquatic environment, while the quantity in water was nearly negligible. In such cases the mmBAF ratio can be approximated as the ratio of the BSAFs of the target chemical and benchmark. However, in contrast to the BSAF, the mmBAF implies no direct mechanistic link between the levels in the sediment and the levels in the organism; sediment need not be the exposure medium of relevance. Rather, the mmBAF expresses the transfer of chemical to the organism from the environment as a whole. A total of n = 19 Bratio values were obtained for ragworms and n = 13 for flounder. The mean of log Bratio and its standard error are plotted in Figure 2. Recall that the extraction of the cVMS from the biota samples was incomplete (from on average 84% for D4 in ragworms to 56% for D6 in flounder)14 so the Bratio values will tend to be underestimated. To preclude an overestimation of the Bratio for D4, only Bratio values were included where CCorg was above the LOQ (n = 8 for ragworms and n = 4 for flounder; note that measured values below the LOQ would be significantly overestimated due to the blank contribution). Since CCsed of D4 was below the LOQ in all samples (i.e., CCsed was overestimated), the Bratio values for D4 in Figure 2 are likely underestimated.
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Figure 3. Schematic explanation of the bioaccumulation of D5 and PCB 180, assuming that both chemicals have the same concentration in sediment. The resulting concentrations in water (freely dissolved, assumed to be in equilibrium with sediment, determined by KOC), trophic level 1 (TL1, lipid normalized, assumed to be in equilibrium with water, determined by KOW), and trophic levels 2 and 3 [TL2 and TL3, lipid normalized, a consequence of biomagnification (PCBs) of biodilution (D5)] are shown.
Similarly, the CCorg of D6 was below the LOQ in all flounder samples but one, so the Bratio values for D6 in flounder in Figure 2 should rather be viewed as an upper boundary. Note that similar results were obtained when PCB 138 and PCB 153 were used as benchmark chemicals (see Figure S1). D5. The mean log Bratio of D5 exceeds 0 (i.e., Bratio exceeds 1) in both ragworms and flounder. This indicates that D5 bioaccumulates in these organisms to a greater extent than PCB 180. Given that PCB 180 is known to be strongly bioaccumulative, this suggests that D5 is also strongly bioaccumulative in benthic invertebrates and fish. As noted in the Introduction, the bioaccumulation of D5 is the focus of scientific debate, with a laboratory measurement of a high fish BCF indicating it is bioaccumulative, while a field measurement of TMF
