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Please be aware that this old REACH registration data factsheet is no longer maintained; it remains frozen as of 19th May 2023.

The new ECHA CHEM database has been released by ECHA, and it now contains all REACH registration data. There are more details on the transition of ECHA's published data to ECHA CHEM here.

Diss Factsheets

Environmental fate & pathways

Endpoint summary

Administrative data

Description of key information

Additional information

MTBE is resistant to hydrolysis at environmentally relevant pH values as demonstrated in an OECD guideline, GLP-compliant study. Strong acids can contribute to the hydrolysis of MTBE but the pH needed for decomposition is far below that normally detected in natural soil and water (Lyman et al. 1982).

According to existing data, the degradation half-life of MTBE in the air is 3-6 days depending on environmental conditions (predominantly OH-radical concentration). Using a degradation rate constant of 2.84E-12cm3/molecule/s and an OH-radical concentration of5E05 radicals/cm3 a half-life of 5.65 days is calculated.

Direct photolysis will not be an important removal process since aliphatic ethers do not absorb light at wavelengths >290 nm. The UV-spectrum (max t 289 nm) indicates that direct photolysis in water may not occur.

MTBE is not readily biodegradable in the aquatic environment according to standard aerobic ready-biodegradation tests (Hüls, 1991a; RBM, 1996a; Slovnaft VÚRUP, a. s., 2005a). However, high degradation rates have been observed in non-standard tests using special types of inoculum, pure cultures and mixed cultures (Shell, 1993; Salanitro et al., 1994; Steffan et al., 1997; Cano et al., 1999; Kharoune et al., 2001; 2002). These studies show that at least some microbial species are capable of degrading MTBE and to use it even as their sole carbon source. It may be considered that MTBE is inherently biodegradable under certain conditions in the aquatic aerobic environment. However, the non-standard test data available indicate that MTBE degradation might not fulfil the test criteria (of OECD 302). In contrast, it also shows that adapted sewage sludge is able to rapidly degrade MTBE.

It could be assumed that where there are continuous releases of MTBE to a STP, such as for large production and processing sites, sewage sludge will have become adapted to the substance and in these cases, the substance could be considered as readily biodegradable. For professional and consumer releases and releases on the regional scale, where adaptation may not occur, the non-standard test data available indicate that MTBE degradation might not fulfil the test criteria (of OECD 302), and a slower degradation categorisation as “inherently biodegradable, not fulfilling criteria” could be considered.

This approach is consistent with the conclusion from the EU Risk Assessment Report for MTBE, which concludes “MTBE is inherently biodegradable under certain conditions in aquatic aerobic environment” (European Commission, 2002).

However, owing to the lack of ready biodegradation seen in the available standard screening tests, and as mandated by ECHA via a substance evaluation decision, a conservative approach to the available data means that the conclusion ‘under test conditions no biodegradation observed’ has been used for exposure assessment purposes.

In anaerobic, static sediment/water microcosms, MTBE does not biodegrade (Suflita et al., 1993; Mormile et al., 1994). Under mixed aerobic/anaerobic conditions biodegradation may in some cases be a significant removal process of MTBE in aerobic sediment (Bradley et al., 1999).

Several studies are available for degradation of MTBE in soil. The results are conflicting. In a study in which soil was polluted with gasoline containing MTBE it was shown that aerobic biodegradation was observed after the spill (Yuan, 2006). This behaviour of MTBE was also observed by Borden et al. (1997). However, other studies concluded that rapid and reliable biodegradation of MTBE in soil cannot be assumed under any normal environmental conditions (both aerobic and anaerobic), indicating very slow degradation in soil (Yeh and Novak, 1994; Allard et al., 1996; Reisinger et al., 2000). As the study by Yuan (2006) was better in design and reporting than the other studies mentioned, the worst-case half-life of 101.6 days in soil from this study is used in the assessment.

The rate constants used in the assessment are:

Degradation for hydrolysis

0 d-1

Degradation for photolysis

0 d-1

Degradation in air

0.123 d-1

Degradation in a non-adapted STP

0 d-1

Degradation in an adapted STP

Monod kinetics (default values)

Biodegradation in water

4.62E-03 d-1

Biodegradation in aerated sediment

2.31E-03 d-1

Biodegradation in soil

1E-03 d-1

Monitoring data support the paramaters used in the modelling, as discussed in Chapter 9 on Exposure Assessment.


Whole-body bioconcentration factors (BCF) of 1.5 and 1.4 were reported for Japanese carp exposed to 10 and 80 mg/l MTBE in a flow-through system at 25 ºC. Fish exposed for 28 days and then transferred to clean water eliminated almost all MTBE residues within 3 days (Fujiwara et al., 1984). The BCF indicate a low potential for bioconcentration. The BCF of 1.5 l/kg is used in the assessment.

The organic carbon-water partitioning coefficient (Koc) calculated from the octanol-water partition coefficient (log Kow = 1.06) using the equation from the Technical Guidance Document (2003) (predominantly hydrophobics) is 9.1 l/kg (log value = 0.95). This predicted value is used in the assessment.

The Henry's Law constant (H) is calculated as 69.8 Pa m3/mol (log H = 1.84), based on a vapour pressure of 33 kPa at 25 °C and a water solubility of 42,000 mg/l at 25 °C in EUSES, this corresponds to a Henry's Law constant of 33.3 Pa m3/mol at environmental temperature.

The Level I fugacity model is used to calculate the theoretical distribution of MTBE between four environmental compartments (air, water, soil, sediment) at equilibrium in a unit world. The model calculates that 93.9% of MTBE partitions to the atmosphere.