Registration Dossier

Environmental fate & pathways

Biodegradation in water and sediment: simulation tests

Currently viewing:

Administrative data

Link to relevant study record(s)

Description of key information

Key value for chemical safety assessment

Additional information

Assimilatory nitrate reductases

Assimilationis the conversion nitrate into ammonium for anabolic reactions. Nitrate is reduced for this purpose by enzymes to nitrite (assimilatory nitrate reductases), and then to ammonia. The assimilatory nitrate reductases are molybdenum-containing enzymes, which are widespread in bacteria, fungi, yeasts, and algae (Campbell, 2001; Inokuchi et al, 2002; Joseph-Horne et al, 2001; Siverio, 2002). Nitrate and chlorate are structurally analogous to each other and may potentially be incorporated into the same enzyme active site, as is evidenced by various assimilatory nitrate reductasesof micro-organisms and plants. Chlorate reduction by assimilatory nitrate reductases has been detected in bacteria (Escherichia coli) using a mutant (Motohara et al, 1976). Balch (1987) experimented with36Cl chlorate as a tracer to study nitrate uptake. In Skeletonema costatum and Nitzschia closterium chlorate was transported into the cells. The ability to reduce chlorate in whole cells has been shown in Ankistrodesmus braunii and Chlorella fusca both algae (Rigano, 1970; Tromballa and Broda, 1971). Chlorate is also a substrate for assimilatory nitrate reductase of Chloralla vulgaris (Solomonson and Vennesland, 1972). The assimilatory nitrate reductases convert chlorate to a toxic product chlorite. These results demonstrate that there is a potential for chlorate reduction under aerobic conditions provided that organisms are present capable of utilizing nitrate as nitrogen source.

Dissimilatory nitrate reductases

Denitrification is a process by which bacteria convert nitrate to dinitrogen that is lost to the atmosphere. Denitrifying bacteria use nitrate instead of oxygen in the metabolic processes. Denitrification takes primarily place where oxygen is depleted and where there is ample organic matter to provide energy for bacteria. Two types of dissimilatory nitrate reductases have been found. One of them is coupled to a complete denitrifying pathway (membrane-bound nitrate reductases; nitrate reductase A), and the other is a periplasmic protein whose physiological role seems to be the dissipation of excess reducing power. Periplasmic nitrate reductases, responsible for denitrification under aerobic conditions are specific for nitrate and not capable of reducing chlorate (Berks et al, 1994; McEwan et al, 1987). Chlorate reduction in denitrifying bacteria is primarily due to membrane-bound nitrate reductase (nitrate reductase A) activity (Iobbi et al, 1987; Morpeth and Boxer, 1985). The reduction of nitrate and chlorate in cell-free extracts of nitrate-grown Bacillus cereus was investigated by Hackenthal (1965). Chlorate reduction rates in cell-free extract were approximately twice as high as the nitrate reduction rates. De Groot and Stouthamer (1969) found that Proteus mirabilis formed different reductases including a chlorate reductase (chlorate reductase C). Chlorate reductase purified from Proteus mirabilis could only use chlorate as a substrate (Oltman et al, 1976). It was found that chlorate reductase was produced constitutively while nitrate reductases were produced inductively. However, the chlorate reduction in cell-free extracts of nitrate-grown bacteria is primarily due to membrane-bound nitrate reductases (de Groot and Stouthamer, 1969).

Chlorite is produced from chlorate by denitrifying microorganisms (Quastel et al, 1925; Karki and Kaiser, 1979). It was found that the absorption spectrum of dissimilatory nitrate reductase obtained from Escherichia coli after oxidation by chlorate was different from that of normal oxidized cytochrome. It was assumed that this was due to the oxidative deformation of the haem by the reduction product chlorite (Itagaki et al, 1963). Chlorite formed by denitrifying bacteria is degraded through chemical reactions with reducing agents such as the protein of nitrate reductase. In conclusion, chlorate reduction associated with nitrate-respiring organisms is a cometabolic process. The rate of chlorate reduction by denitrifying bacteria is therefore directly linked to the rate of denitrification.

Growth linked biodegradation (anaerobic)

It is now well-known that bacteria have evolved that can grow by the anaerobic reductive dissimilation of chlorate into innocuous chloride. Bacteria capable of growing with chlorate as electron acceptor are widely spread nature. This has been shown by (per)chlorate reduction with various energy substrates with a number of enrichment cultures (Bryan and Rohlich 1954; van Ginkel et al, 1995; Logan, 1998). The ubiquity of (per)chlorate reducing microorganisms was also shown quantitatively by enumerating the (per)chlorate-reducing bacteria in very diverse environments, including soils, aquatic sediments, sludges, and lagoons. In all of the environments tested, the acetate-oxidizing (per)chlorate reducing bacteria represented a significant population, whose size ranged from 2.3 × 103to 2.4 × 106cells per g of sample (Coates et al, 1999; Wu et al, 2001). Existence of (per)chlorate respiring bacteria have also been demonstrated in marine waters (Logan et al, 2000)

(Per)chlorate reducing microorganisms are easily enriched and isolated from many environments. All of these organisms could grow anaerobically by coupling complete oxidation of reducing agents to reduction of chlorate at high rates (Table). Under fully aerobic conditions, chlorate is not reduced by capable bacteria. Nitrate can also interfere with chlorate reduction (Chaudhuri et al 2002).

 

Table Growth ratesof various cultures capable of reducing chlorate.

Culture

Reducing agent

Rate (h-1)

Reference

Azospira oryzoa GR-1

Acetate

0.1

Rikken et al (1996)

Dechloromonas Agita sp CKB

 

0.28

Bruce et al (1999)

Azospira sp KJ

Acetate

0.26

Logan et al (2001)

Dechlorosomonas sp PDX

Acetate

0.21

Logan et al (2001)

Dechlorosomonas sp PDX

Lactate

0.15

Logan et al (2001)

PDA

Acetate

0.18

Logan et al (2001)

PDB

Actate

0.21

Logan et al (2001)

Mixed culture

Acetate

0.56

Logan et al (1998)

Mixed culture

glucose glutamate

0.12

Logan et al (1998)

Mixed culture

Phenol

0.04

Logan et al (1998)

 

Azospira oryzae strain GR-1 (DSM 11199) isolated from activated sludge was the first bacterium studied in more detail (Rikken et al, 1996; Wolterink et al, 2005). When strain GR-1 was grown on acetate, the release of chloride was proportional to the disappearance of chlorate, showing that this compound was completely reduced. The oxidation of acetate is coupled to the reduction of chlorate, whereas chlorite reduction is not affected by the addition of acetate. Azospira oryzae strain GR-1 disproportionates chlorite into molecular oxygen and chloride. For chlorate reduction by Azospira oryzae strain GR-1 the following biodegradation pathway was formulated: ClO3-  ClO2- Cl-+ O2. The rapid dismutation of chlorite into chloride and molecular oxygen is the key reaction in the reduction of chlorate. All (per)chlorate-reducing bacteria isolated to date have the ability to dismutate chlorite (Coates and Achenbach, 2004). Complete reduction of chlorate into chloride and molecular oxygen is catalysed by two enzymes. Chlorate is reduced to chlorite by (per)chlorate reductase (EC 1.97.1.1) (Kengen et al, 1999). Chlorate respiration at high rates is made possible by the action of the second enzyme which reduces the toxic chlorite to chloride while producing molecular oxygen. This is mediated by chlorite dismutase (EC 1.13.11.49) (van Ginkel et al 1996; Stenklo et al 2001). Chlorite has never been found to accumulate in solution during bacterial respiration of (per)chlorate.

Aqueous fresh systems

Although sodium chlorate is considered readily biodegradable, the TGD default value does probably not represent the half-live in aerobic surface water. The fate of sodium chlorate in aerobic surface waters is linked to nitrate assimilation rates (cometabolic process).It is therefore important to know the nitrate assimilation rates in any aqueous ecosystems for assessment of the potential of chlorate reduction.An estimate of the nitrate uptake rate is therefore based on data found in the open literature.Compared to nitrate-nitrogen uptake rates, chlorate reduction rates are assumed to be 10 times lower (expert opinion). The total amount of chlorate in the environment compared to nitrate assimilated is probably negligible. Chlorate can therefore be completely transformed. Nitrate-nitrogen uptake rates by phytoplankton in aranged from 0.006 to 0.036µM h-1 (Rojo et al 2008). The nitrate-nitrogen uptake rates by heterotrophic bacteria in the riverrange from 0.001 to 1.44 µM h-1 (Middelburg and Nieuwenhuizen, 2000). Using an environmental chlorate concentration of 0.036 µM – 0.52 µM (Versteegh-Neele-Cleven,1993) half lives ranging from 0.2 day to 18 days can be calculated with the nitrogen uptake rates for phytoplankton (zero order kinetics).

Freshwater with sediment

Reduction of chlorate was found in a water-sediment system under both aerobic and anaerobic conditions (OECD TG 308). Sediments with a high (13% for aerobic sediment and 4.4% for anaerobic sediment) and a low (0.4% both aerobic and anaerobic) organic carbon contents were used. Chlorate was reduced at higher rates in sediments with a high organic carbon content compared to sediments with a low carbon content. This result is consistent with the literature on microbial chlorate reduction. The DT50 for sediment with high organic carbon content was 8 days in water phase and less than 3 days in sediment under aerobic conditions and 9 days in water phase and less than 1 day in sediment under anaerobic conditions. The DT50 for sediment with low organic carbon content was 20 days in water phase and 18 days in sediment under aerobic conditions and 29 days in water phase and 24 days in sediment under anaerobic conditions (van der Togt and van Ginkel, 2005)

Seawater

Biological transformation of chlorate through assimilatory nitrate reductases is probably the only significant sink for chlorate in the marine environment.The fate of sodium chlorate in seawater is therefore primarily linked to nitrate assimilation rates. In thethe nitrate-nitrogen assimilation rate were <0.2 nM day-1 in the oligotropic period and >2.3 nM day-1 during blooms (Lipchultz, 2001).  Compared to nitrate-nitrogen uptake rates, chlorate reduction rates are assumed to be 10 times lower (expert opinion). The total amount of chlorate in the environment compared to nitrate assimilated is negligible. Chlorate can therefore be completely transformed. Lander et al (1994) measured chlorate concentrations of <24 nM in the Baltic sea at 3-4 km from the effluent discharge of a pulp and paper plant using chlorate.

The environmental chlorate concentration in the oceans is expected to be at least 10 times lower than found by Lander et al. (1994). Using this environmental chlorate concentration for seawater of 2.4 nM, half lives ranging from 5.2 days to 60 days can be calculated (zero order kinetics).

Waste water treatment systems

More than 99% chlorate removal was achieved in a continuously fed up-flow fixed bed reactor operated at a hydraulic retention time of only 3.6 hours. Chlorate was reduced with molasse (Detaille et al 1992). Anaerobic fixed-film processes show that bacteria can remove chlorates from kraft bleach effluent with less than one hour retention time. (Malmqvist and Welander, 1992). Complete biological chlorate removal was also achieved in a pilot-scale bioreactor with suspended carrier material at hydraulic retention times as short as 24 minutes (Malmqvist and Welander, 1994). Pilot studies were carried out with anaerobic (to remove chlorate) and aerobic (to oxidize organic matter) reactors in series. The chlorate concentrations in softwood and hardwood effluents averaged 80 (30-180) and 100 (30-160) mg/L, respectively. The COD concentrations ranged from 1.3 to 2.5 and 0.7 to 1.9 g/L for softwood and hardwood effluents, respectively. In this pilot study chlorate removal percentages of >90 were obtained (Malmqvist and Welander, 1993).   

Chlorate levels of 60 to 70 mg/L were reduced to less than 2 mg/L within a two-hours anaerobic pre-treatment period and eight hours of aerobic treatment in a laboratory scale activated sludge system. A denitrifying culture can not remove chlorate under these conditions (Malyk, 1992).

Laboratory anaerobic and aerobic reactors treating kraft bleach plant effluents operated continuously, removed chlorate easily. Removal of chlorate in softwood and hardwood effluents was 99% and 96%, respectively, with little difference in efficiency between the single-stage and two-stage anaerobic systems (Dorica and Elliot, 1994)