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Environmental fate & pathways

Endpoint summary

Administrative data

Description of key information

Additional information

2,4-Dinitrophenol's (2,4-DNP) production and its use in the manufacture of dyes are the main sources of release in the environment through various loss and waste streams. Below some specific releases to several environmental compartments are mentioned, followed by a more in depth analysis of the fate of 2,4-DNP in each specific compartment.

Dinitrophenols are released to the air from manufacturing plants and facilities that use them for the production of explosives, dyes, wood preservatives, pesticides. (HSDB 1994; TRI92 1994).

Sources of 2,4-DNP in atmosphere could be automobile exhaust dusts, pesticides, hazardous waste disposal sites and photochemical reactions of benzene with nitrogen oxides (Nojima et al. 1983) that thus can be considered a secondary source of environmental exposure.

Entry into surface water can occur because of releases from the above mentioned production facilities and by industries that manufacture other products using 2,4-DNP.

2,4-DNP has for instance been detected in the waste water from nitrobenzene-manufacturing plants (Patil and Shinde 1988), industries or plants that are explosives producers (McLuckey et al. 1985), dye-manufacturing plants (Games and Hites 1977), and sewage treatment plants where the influent waters already contain dinitrophenols (DeWalle et al. 1982). Since 2,4-DNP is also used to produce picric acid, picramide, and photographic developer (diaminophenol hydrochloride) and in preserving wood (Hawley 1981), waste waters or land runoffs from these industries may also release 2,4-DNP to surface water.

Dinitrophenols may be released in soils in the vicinity of the sites where they are manufactured and used by wet and dry deposition or from waste water release. Dinitrophenols were found in the soil of a decommissioned wood preserving facility (EPA 1988a).


Organic substances with vapour pressures of 10^-4 to 10^-8 mmHg at ambient temperature should exist partly in the vapour and partly in the particulate phase in the atmosphere (Eisenreich et al. 1981). According to a model of gas/particle partitioning of semi-volatile organic compounds in the atmosphere, 2,4-DNP, which has a vapour pressure of 1.49 X10^-5 mm Hg at 18 °C, is expected to exist as a vapour in the ambient atmosphere.

Studies indicate that photo-degradation may be an important fate process, although the kinetic of these reactions are unknown (Atkinson et al., 1992). 

Vapour-phase 2,4-DNP is degraded in the atmosphere by direct photolysis by sunlight at > 290 nm (HSDB, Sadtler Res lab, 2012) and by the reaction with photochemically-produced hydroxyl radicals (SRC) with an half-life estimated of 28 days (degradation rate constant OH radicals of 5.76 x 10^-13 cm3).

Conversely, reaction with nitrate radicals should not be a significant environmental fate (by structural analogy) (Grosjean D. 1985).

The presence of dinitrophenols in air is < 100%. This is explained in a study of Nojima K, Kawaguchi A, Ohya T, et al. 1983 where the detection of 2,4-DNP by GC in air are indirect indications of the partially sorbing in particulate matter.

Hence, since dinitrophenols are expected to be present partly in the particulate phase in the air, the reaction rate in air is expected to be even slower than the estimated value for the gas phase reaction (Atkinson R. et al., 1988), so significant removal of dinitrophenols from the atmosphere due to photochemical or other chemical reactions is not likely. 

In conclusion, DNP exists as a vapour in the atmosphere. Direct and indirect photodegradation can occur, however the main elimination pathway seems to be sorption to particulate matter in air with subsequent wet and dry deposition of DNP. The presence of DNP in rain and snow was experimentally verified by Alber et al. 1989; Cape1 et al. 1991; Levsen et al. 1990). In fact, the detection of 2,4-DNP in rain, and snow showed that at least partial removal of these compounds occurs by physical processes (Alber M, Bîhm HB, Brodesser J, et al. 1989).

WATER COMPARTMENT (incl. sediment)

Neither photochemical nor other chemical processes have been identified as significant for the transformation/degradation of dinitrophenols in natural waters (Callahan et al. 1979; Lipczynska-Kochany 1992; Tratnyek and Hoigne 1991; Tratnyek et al. 1991). Furthermore, the direct photolysis of 2,4-DNP in water is too slow to be an important environmental fate process (Lipczynska-Kochany E, 1991).

This conclusion has been explained in several studies (Tratnyek et al. 1991- Mabey WR, et al. 1981), where the main conclusion was that the photo transformation of 2,4-DNP would not be important in water (half-life of ≈500 days). This value is based on the reaction between 2,4-DNP, singlet oxygen (O2) and peroxy radicals (RO2) concentrations (the rate constant of 2,4-DNP’s reaction with singlet oxygen concentrations is 4.05 x 10^5molar-second and the estimated average of the singlet oxygen concentration in typical eutrophic fresh water is 4 x 10^-14molar).

Usually, concentrations of singlet oxygen and peroxy radicals in typical eutrophic waters are low (10^-12 and 10^-9 molar, respectively). Hence, the reaction of hydroperoxy radicals (HO2) with 2,4-DNP, producing a ring hydroxylated products, would not be significant (Mill and Mabey 1985).

2,4-DNP may be also photoreduced to 2-amino-4-nitrophenol in the presence of ascorbic acid or ferrous ions, and the reaction is sensitized by chlorophyll. The possibility of such photo reduction exists in natural water in which the suspended reducing matter may act as a reducing agent and humic substances or algae may serve as a sensitizer (Massini P, Voorn G. 1967).

2,4-DNP is not expected to undergo hydrolysis in the environment due to the lack of functional groups that hydrolyse under environmental conditions (Lyman WJ et al., 1990).

The sorption and subsequent transport of dinitrophenols from water to suspended solids and sediment would be significant in natural waters that are acidic and/or have high organic matter and clay content. In studies (Kaufman, 1976 and Callahan et al., 1979) it has been indicated that the sorption of dinitrophenols by soil or sediment would depend on their organic carbon content, clay content, and pH.

An increase in clay and organic carbon content and a decrease in pH would increase the amount sorbed.

Biodegradation may be the most important process of loss for dinitrophenols in natural waters. 

Excepting results from a Japanese MITI test biodegradation, where the 2,4-DNP reached 0% of its theoretical BOD in 4 weeks using an activated sludge inoculum (aerobic biodegradation) (NITE; Chemical Risk Information Platform (CHRIP), in general, complete or partial biodegradation of 2,4-DNP was observed in several aquatic systems.

The half-life in aerobic waters for biodegradation has been reported to be 68 days, in anaerobic water 2.8 days (Capel PD, Larson SJ, 1995).

Considering different microbial systems, under aerobic conditions biodegradation was observed in mixed microorganisms from activated sludge (Kincannon et al. 1983a, 1983b; Patil and Shinde 1989; Pitter 1976), enriched sewage (Brown et al. 1990; Wiggins and Alexander 19SS), adapted sediment from rivers or waste lagoons (Barth and Bunch 1979; Chambers et al. 1963; Tabak et al. 1964).

In several studies with activated sludge previously adapted to mineralize low concentrations of dinitrophenols (Jo KW, Silverstein J, 1998) and under methanogenic conditions with anaerobic digester sludge (Battersby and Wilson 1989; O’Connor and Young 1989) biodegradation was observed, but this activity diminished at higher concentrations of 2,4-DNP.

Toxic effects were explained in the IUCLID section 6.1.7 (ecotoxicological information for microorganisms).

In fact toxic level of c.a 10 -20 mg/L of 2,4-dinitrophenol was defined (O'Connor and Young, 1989) and a detailed study on sludge activity indicates a NOEC (no inhibitory effect) of 4 ppm.

2,4-DNP is also biodegraded by several pure cultures of microorganisms. Usually, the pure cultures are able to biodegrade 2,4-DNP after a certain adaptation period and as long as the concentration of 2,4-DNP is below a certain toxic level. The degradation pathway depends on the microorganisms and the conditions of aeration.

Typically, with aerobic organisms and aerobic conditions, the biodegradation proceded by replacement of nitro groups by hydroxyl groups and liberation of nitrite, or by hydroxylation of the aromatic ring positions 3, 5, or 6 (Raymond and Alexander 1971). 

Although pure culture studies are important for establishing degradative pathways, they do not reflect real environmental situations where mixed microorganisms and different nutritional conditions are present.

Complete or partial biodegradation of 2,4-DNP was observed also in field conditions, as in an aeration lagoons and settling ponds.

In water compartment for 2,4-DNP, estimated and experimental BCF values are available.

The meaning of BCF values indicates the possibility of bioconcentration in the aquatic compartment. 

The overall range of BCF measured in salt water is from 3.0 to 16 (EPA databank). Additional studies measured BCF values of < 0.4 - 0.7 and < 3.7.

These BCFs suggest that the potential for bioconcentration in aquatic organisms for 2,4-DNP is low (SRC) (NITE; Chemical Risk Information Platform (CHRIP)). Estimated BCF has a value of 0.56. The concentration of 2,4-DNP in fish may be even lower than its concentration in water. (McCarty LS, Mackay D, Smith AD, et al. 1993). Furthermore, bioconcentration of dinitrophenols from water to aquatic organisms and from soil to plants is not expected to be important.

No data were located on the biomagnification potential for dinitrophenols in predators that consume contaminated prey (EPA 1986a; O’Connor et al. 1990).


It has been speculated that 2,4-DNP in soil may be reduced to 2-amino-4-nitrophenol by sunlight in the presence of a reductant, such as ferrous ions and a sensitizer, such as chlorophyll (Kaufman 1976; Overcash et al. 1982; Shea et al. 1983). However it seems unlikely that sunlight would penetrate the soil surface layer.

Koc values collected suggest that 2,4-DNP is expected to have high mobility in soil (Martins JM, Mermoud A: J Contam Hydrol, 1998). Measured Koc values are in the range of 13.5 - 16.6 and estimated value is 284.3 L/kg (Kow method). Moreover the pKa of 2,4-DNP is 4.09, which indicates that this compound will exist primarily as an anion in moist soil surfaces and anions are expected to have very high mobility in soils (SRC).

Volatilization of 2,4-DNP from moist soil surfaces is not expected to be an important fate process (SRC) since the anion will not volatilize. The neutral species has a Henry's Law constant of 8.71E-003 Pa-m3/mole at 20°C and a vapour pressure of 1.49 X10-5mm Hg at 18 °C (Wild and Jones (1992). The possibility of volatilization of phenolic compounds in soil via co-distillation with water has been suggested and release to air via aerosol formation is possible (Kincannon and Lin, 1985).

The mobility of dinitrophenols in soils decreases with increase in acidity, clay, and organic matter content. The ionized form is more water soluble and moves faster through soil. The transport of dinitrophenols from soil to groundwater may also occur via leaching. The amount of DNP leached depends on the dinitrophenol adsorption capability of soils. 2,4-DNP is a moderate weak acid that is expected to be highly labile (leachable and plant available) in higher-pH soils. Several studies defined that adsorption of phenols in soil increases with a decrease in pH and an increase in organic carbon, goethite (one of the most common iron oxides in soil) and clay content (Hudson-Baruth and Seitz 1986; Kaufman 1976; O’Connor et al. 1990; Shea et al. 1983; Stone et al. 1993).

Biodegradation may be the most significant process for destroying dinitrophenols in soil.

The biodegradation half-life of 2,4-DNP in an acidic soil was reported of 32.1 days and the biodegradation half-life in a basic soil as 4.6 days (Loehr RC, 1989).

Depending on the type of soil (pH, organic matter content), the length of acclimation phase, as well as the initial concentration, the residence time of dinitrophenols for the aerobic biodegradation of soil may vary from 8 to 120 days (Kincannon and Lin 1985; Loehr 1989; O’Connor et al. 1990).

2,4-DNP can be biodegraded by isolated culture proceeding of the reduction of the nitro group or displacement of a nitro group by a hydroxyl group with the release of nitrite ions (Kohping GW, Wiegel J. 1987, Shea PJ, Weber JB, Overcash MR., 1983). However, high concentration 2,4-DNP (see IUCLID section 6.1.7) may be toxic to the degrader microorganisms.

The biodegradation of dinitrophenols in soils will occur also by bacteria in multiphasic mineralization kinetics (involving several slow types of anaerobic reaction anaerobic to methane and carbon dioxide) (Schmidt SK, Gier MJ. 1990, Young LY. 1986). Pure culture of the fungus Fusarium oxysporum, pure cultures of Nocardia alba, Arthrobacter and Corynebacterium simplex are all able to reduce 2,4-DNP (Overcash MR et al, 1982).

Some loss of dinitrophenols from soil could occur by plant uptake. Bioaccumulation factor (concentration in plant over concentration in soil) in lettuce, carrot (tops, peels, and root), hot pepper foliage, and fruits is < 0.01 at a soil pH of 6.7-7.2.

Since dinitrophenols undergo metabolism in plants, plants accumulation of dinitrophenols due to uptake would not be significant. Since the concentration of the non-ionized form is only <0.25% of total DNP at pH 6.7, the soil pH has to be considerably lower to permit the uptake of the non-ionized form in a significant way in plants (O’Connor GA, Lujan JR, Jin Y. 1990).

In fact, in a Shea PJ, Weber JB, Overcash MR., 1983 study it has been demonstrated that the uptake and the translocation could be significant in soil with low pH where the concentration of non-ionized dinitrophenols (more readily adsorbed than the ionized form) are already higher.