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Survey of azo-colorants in Denmark

Toxicity and Fate of Azo Dyes

Physico-chemical properties

The molecular weight for the azo dyes included in the present survey lies within the range of 197 to 996 g/mol. The ranges and mean values for the different chemical classes are listed in Table 5.1.

Table .1

Molar weight for azo dyes used in Denmark.

Molvægt for azofarvestoffer anvendt i Danmark.

 

Number of observations

Mean
g/mol

Range
g/mol

Acid

11

582

351-830

Basic

2

-

248-260

Direct

8

807

698-996

Disperse

1

625

-

Mordant

n.o.1

n.o.1

n.o.1

Solvent

8

273

197-379

Reactive

n.o.1

n.o.1

n.o.1

1 No observations.

General aspects

As described in chapter 3, section 3.2 the dyes may be divided into water soluble cationic and anionic dyes and water insoluble dyes - non-ionic dyes.

The basic dyes are cationic. The acidic, direct and reactive, dyes are anionic. The disperse, mordant and solvent dyes have a low water solubility. These dyes are basically characterised as non-ionic or neutral dyes, and thereby hydrophobic in character.

The electron-withdrawal character of azo-groups generates electron deficiency. Thus it makes the compounds less susceptible to oxidative catabolism, and as a consequence many of these chemicals tend to persist under aerobic environmental conditions (Knackmuss, 1996)

Furthermore, dyes must have a high degree of chemical and photolytic stability in order to be useful. It is thus unlikely that they, in general, will give positive results in short-term tests for aerobic biodegradability (e.g. OECD), (Brown & Anliker, 1988). Stability against microbial attack is also a required feature of azo dyes (Pagga & Brown, 1986), because it may prolong the lifetime of the products, in which azo dyes are applicable.

Subsequently, photolysis is not considered to be an important degradation pathway for azo dyes. Even though, all the azo dyes have absorption maxima in the range of visible and UV-light.

Vapour pressure data are not available for most of the azo dyes. In Table 5.2, a few examples are listed. They clearly indicate that the vapour pressure, in general, is very low.

Table .2
Examples of vapour pressures.
Eksempler på damptryk.

Compound

Vapour pressure (mmHg)2

Acid Yellow 10

2,5x10-20

Solvent Yellow 2*1

3.6x10-8

Disperse Red 9

1.9x10-11

Disperse Red 1

2.3x10-13

Ref.: Baughman & Perenich (1988b).
1 See footnote.
2 It is not stated, if the vapour pressure is measured or estimated.

Ionic azo dyes

In general the ionic azo dyes will be almost completely or partly dissociated in an aqueous solution. Solubility in the range 100 mg/l to 80,000 mg/l has been reported for the ionic azo dyes ( HSDB, 1998). In addition, they would be expected to have a high to a moderate mobility in soil, sediment and particular matter, indicated by the low Koc values. However, due to their ionic nature, they adsorb as a result of ion-exchange processes.

In addition, ionic compounds are not considered to be able to volatilize neither from moist nor dry surfaces, and the vapour pressures for these dyes are very low, e.g. Acid Yellow 10.

Only the reactive dyes show a high degree of hydrolysation. Reactive dyes form covalent bonds to the textile. The fixation competes with the reaction of the leaving group of the reactive dye with water (hydrolysis). Therefore the non-fixed dye in a dye bath is the hydrolysed derivative, which has no more the characteristics of the reactive substance. One of the characteristics of these reactive dyes, with a few exceptions, is that the aromatic moieties carry sulfonic groups. Chemical or enzymatic reduction leads to the formation of amino sulfonic acids (ETAD, 1991).

Estimated Kow values for the ionic dyes are generally very low, e.g. 2.75 x 10-5 for Acid Orange 10* and 100 for Direct Black 38*.

Non-ionic azo dyes

The solubility in water is in the range of 0.2 mg/l to 34.3 mg/l for the solvent dyes included in the present survey (HSDB, 1998; Baughman & Perenich, 1988a).

As stated above, vapour pressures are not available for most of the azo dyes, but they are generally low, as shown in Table 5.2. However, some of the disperse dyes have vapour pressures high enough for application from the vapour phase. Furthermore, disperse dyes are believed to dye fabrics by the same mechanism by which hydrophobic pollutants adsorb onto sediments, and the equilibrium can be described by a partition coefficient (Baughman & Perenich, 1988b).

Disperse dyes are the main group of hydrophobic dyes, thus they have a significant potential to adsorb sediments and bioconcentrate (Yen et al., 1991). Disperse dyes are further more highly lipophilic (Anliker, 1986). Solvent dyes are, like disperse dyes, neutral hydrophobic dyes (Baughman & Perenich, 1988b).

The solvent dyes are large, complex molecules, that can be expected to have lower vapour pressures than disperse dyes (Baughman & Perenich, 1988b).

The partition coefficients (Kow) are very high for the non-ionic dyes. In the range of 420 for Solvent Yellow 1* to 11,220 for Solvent Yellow 2. The disperse dye Disperse Blue 79* has a Kow of 3,630. The values are all based on estimates.

Metabolites

Generally, the physico-chemical parameters vary within the following 4 groups: aniline, toluidine, benzidine and naphthalene. These are potenti- ally carcinogenic aromatic amines, which are among the cleavage products and impurities of the azo dyes.

The solubility in water varies. Some are almost insoluble (e.g.

4,4´-methylenebis [2-chloroaniline] and 3,3´-dimethoxybenzidine), whereas others are highly soluble, up to 16.8 g/l (o-toluidine).

The absorption maxima are generally in the range of 240 to 300 nm, i.e. below the range of visible and UV-light.

The vapour pressures are in the range of 7.5 ´ 10-7 to 0.32 mmHg.

The estimated partition coefficients (Kow) lay within the range of 21 for benzidine to 8,300 for 4-o-tolylazo-o-toluidine.

Summary

The azo dyes may be subdivided into two groups: the ionic and non-ionic dyes. They have some common features, though. Their absorption maxima is in the range of visible and UV-light and the vapour pressures, if available, are very low in the range of 2.5 ´ 10-20 to 3.6 ´ 10-8 mmHg. The hydrolysation is, except for the reactive dyes, very low.

However, the two groups also exhibit major differences. In general, the ionic azo dyes will be almost completely or partly dissociated in an aqueous solution. The non-ionic dyes, on the other hand, are only sparingly soluble (<100 mg/l). The estimated Kow values for the ionic dyes are generally very low e.g. -2.75 ´ 10-5 for Acid Orange 10* and 100 for Direct Black 38*. However, the non-ionic dyes have very high partition coefficients (Kow ), e.g. 3,630 Disperse Blue 79* and 11,220 for Solvent Yellow 2.

The solubility of the metabolites varies similarly from almost insoluble to very soluble. The absorption maxima are generally below the range of 240 to 300 nm. The vapour pressures are in the range of 7.5 ´ 10-7 to 0.32 mmHg.

The estimated partition coefficients (Kow) lay within the range of 21 for benzidine to 8,300 for 4-o-tolylazo-o-toluidine.

Toxicity

Acute toxicity

The acute toxicity of azo dyes, as defined by the EU criteria for classification of dangerous substances, is rather low. Information about acute oral toxicity, including skin and eye irritation, is in form of material safety data sheets available for many commercial azo dyes. Only a few azo dyes showed LD50 values below 250 mg/kg body weight, whereas a majority showed LD50 values between 250-2,000 mg/kg body weight (Clarke & Anliker, 1980). Remazol Black Bâ (Reactive Black 5) represents an important group of newer azo dyes, namely the reactive dyes. For this dye a comprehensive study on acute toxicity was carried out. The study showed that LD50 exceed 14,000 mg/kg body weight, and that the dye was neither irritant to skin nor to eye (Hunger & Jung, 1991).

Exposure to aromatic amines may cause methemoglobinemia. The amines oxidise the heme iron of haemoglobin from Fe(II) to Fe(III), blocking the oxygen binding. This results in characteristic symptoms like cyanosis of lips and nose, weakness and dizziness. The extent of which various aromatic amines can cause methemoglobinemia varies, however, widely (Ullmann, 5th Edition).

Sensitisation

Occupational sensitisation to azo dyes has been seen in the textile industry since 1930. The first observations were made in 1930 when 20% of the workers dyeing cotton with red azoic dyes, developed occupational eczema (Foussereau et al., 1982).

Attributing an allergy to a particular azo dye is a complex and difficult process, due to the following reasons:
a great number of azo dyes, approximately 2,000.
each azo dye is marketed under several different names.
azo dyes very often contain impurities.

This may be the reason why, in rather rare cases, exposure to azo dyes has led to recognition of a possible relationship between skin sensitisation and a particular azo dye.

The majority of sensitising dyes, present in clothes, practically all belong to the group of disperse dyes, which has been developed for use on synthetic fibres. The explanation is probably that the attachment of molecules from disperse dyes is weak, as they are more easily available for skin contact.

In clinical patch tests the following azo dyes have shown sensitising properties (Cronin, 1980):
Disperse Red 1, 17.
Disperse Orange 1, 3, 76.
Disperse Yellow 3, 4.
Disperse Blue 124.
Disperse Black 1, 2.

In Germany, disperse azo dyes like Disperse Blue 1, 35, 106 and 124, Disperse Yellow 3, Disperse Orange 3, 37, 76 and Disperse Red 1 have been associated with contact dermatitis, resulting from exposure to textiles coloured with these dyes. In most cases the dermatitis resolved, once the sensitising "textile" had been discarded. These dyes are no longer recommended for colouring of textiles, which come into contact with the skin (Platzek, 1995).

Non disperse azo dyes, used for colouring of natural fibres were investigated in 1,814 patients attending the clinic patch tests (Seidenari et al, 1995). 0.88% of the patients reacted positive to the following dyes: Direct Orange 34 (8 patients), Acid Yellow 61 (5 patients), Acid Red 359 (2 patients) and Acid Red 118 (1 patient).

Remazol Black Bâ (Reactive Black 5) was investigated for sensitisation potential in experimental animals and was found to be negative. However, a few cases of allergic reactions have been observed in man.

Despite a very broad application field and exposure, sensitising azo dyes have been identified in relatively few reports (Cronin, 1980).

Toxicokinetic

Only limited information is available regarding absorption, distribution, and excretion of azo dyes, whereas the metabolism after administration of oral consumption has been investigated extensively. Absorption of azo dyes through the skin is doubtful, as intact azo dyes may not penetrate the skin (NIOSH, 1980).

A distribution study conducted with a 14C-biphenyl ring, labelled Direct Blue 15* and Direct Red 2, in rats showed that liver, kidney and lung accumulated and retained higher levels of 14C than other tissues, 72 hours after administration of a single oral consumption (HSELINE, 1998).

The azo linkage is the most labile portion of an azo dye molecule and may easily undergo enzymatic breakdown in mammalian organisms, including man. The azo linkage may be reduced and cleaved, resulting in the splitting of the molecule in two parts (Brown & DeVito, 1993).

The anaerobic environment of the lower gastrointestinal tract of mammals is well suited for azo-reduction. Several anaerobic intestinal bacteria are capable of reducing the azo linkage. The majority of these bacteria belong to the genera Clostridium and Eubacterium. They contain an enzyme associated with the cytochrome P 450, also termed azo-reductase. It is a non-specific enzyme, found in various micro-organisms and in all tested mammals (NIOSH, 1980).

In mammalian organisms azo-reductases are, with different activities, present in various organs like liver, kidney, lung, heart, brain, spleen and muscle tissues. The azo-reductase of the liver, followed by the azo-reductase of the kidneys possess the greatest enzymatic activity.

Although reduction and cleavage of the azo-linkage is the major metabolic pathway of azo dyes in mammals, other metabolic pathways may take place. Major routes of detoxifying metabolism of azo dyes and aromatic amines are ring hydroxylation and glucuronide conjugation.

After cleavage of the azo-linkage, the component aromatic amines are absorbed in the intestine and excreted in the urine (Brown & DeVito, 1993). However, the polarity of azo dyes influences the metabolism and consequently the excretion. Sulphonation of azo dyes appears to decrease toxicity by enhancing urinary excretion of the dye and its metabolites. Sulphonated dyes, mainly mono-, di- and trisulphonated compounds are world-wide permitted for use in foods, cosmetics and as drugs for oral application.

Highly sulphonated azo dyes are poorly absorbed from the intestine after oral intake. Practically a complete cleavage of the azo linkage takes place in the gastrointestinal tract. This results in sulphonic acids rather than aromatic amines. These acids are rapidly absorbed, modified by the liver and excreted in the bile and urine. Sulphonated, fat soluble azo dyes are not reduced by the gut micro-organism but absorbed from the intestine and metabolised to the more polar N- or O-glucuronide and excreted as glucuronide conjugates (Parkinson & Brown, 1981).

The aromatic component amines of azo dyes may be absorbed into the body through the lungs, the gastrointestinal tract or the skin (ECDIN, 1993).

Mutagenicity

In general, the correlation between results of mutagenicity tests and carcinogenicity shown in animal experiments of azo dyes, is poor. The lack of correlation is probably due to the rather complex metabolic pathways, which azo dyes undergo in mammalian organisms (Brown & DeVito, 1993).

The majority of azo dyes requires metabolic activation, namely reduction and cleavage of the azo linkage to the component aromatic amines to show mutagenicity in vitro test systems. Therefore the majority of azo dyes, if highly purified, will, at least without metabolic activation, be negative in such tests (Arcos & Argus, 1994).

Many of the commercial available azo dyes may, however, due to impurities, e.g. contamination with aromatic amines, show mutagenic activity in vitro.

Carcinogenicity

Since the mid-nineteenth century the growth of the synthetic dye industry and in particular the azo dye industry has been based on aromatic amines and consequently contributed to a serious occupational exposure.

Correlation between exposure of aromatic amines and human cancer was reported as early as 1895 by Rehn. He reported four cases of bladder cancer, named as "aniline cancer", out of several hundreds of workers engaged in the manufacture of fuchsin from crude aromatic amines for 15-29 years.

Between 1921 and 1951 Case computed a number of bladder cancer deaths for men manufacturing azo dyes and compared this to the expected incidence of bladder cancer in England. Four bladder cancer deaths were expected, whereas 127 deaths were found. Approximately 25% of all workers being exposed to aromatic amines, including 2-naphthyl- amine and benzidine, developed bladder cancer. The workers, who were only exposed to benzidine, had fewer tumours (15%) than those being exposed to 2-naphthylamine (50%). A few workers, who distilled 2-naphthylamine, all died of bladder cancer (Cartwright, 1983).

Besides the historical evidence, case-control studies have later been carried out on several occupational groups, including machinists, cooks, hairdressers, coal miners, carpenters etc. In several occupational groups a low to an elevated risk of bladder cancer was seen (Miller & Miller, 1983).

For decades there has been a strong human evidence for the association of bladder and renal pelvis cancers with specific aromatic amines. In addition, there has been an evidence, although weaker, that stomach and lung cancers are also associated with exposure to these amines. Aromatic amines do not induce tumours in humans at the exposure site, e.g. lungs and skin, but usually at a site as the urinary bladder.

The latency period, namely the period between the first exposure and the diagnosis of bladder cancer, ranged from 5 to 63 years. The average latency period was approximately 20 years, but cases of cancer after a few months of exposure have also been described (Cartwright, 1983).

Association between aromatic amines and bladder cancer in humans lead to extensive examination of the possibility for induction of bladder cancer in experimental animals.

In experimental animals, aromatic amines induced tumours in liver, intestine or urinary bladder. Furthermore, tumours in mammary gland and the skin were observed in rats (Sontag, 1981).

The carcinogenicity of aromatic amines is species specific. In experimental animals, benzidine was carcinogenic after administration of oral consumption and subcutaneous injections, producing liver tumours in rats, mice and hamsters, whereas bladder cancers were only seen in dogs.

2-Naphthylamine was a potent bladder carcinogen in dogs, but it was non-carcinogenic in rats and rabbits. After treatment with substituted benzenediamine, the incidence of bladder cancer in treated rats was only slightly elevated, but in addition, kidney tumours were observed (Clayson & Garner, 1976).

Although the latency period for human bladder cancer is relatively long, this period may be very short for animal carcinogenesis. Dyes based on benzidine, namely Direct Black 38*, Direct Blue 6* and Direct Brown 95* were investigated in a 13 week subchronic feeding study in rats. All these dyes induced a high incidence of pathological changes (neoplastic nodules) and/or liver cancer within 5 weeks. This is most probably the shortest latency period known for any chemical study with carcinogenic properties (Clayson & Garner, 1976).

Molecular mechanism of carcinogenicity

There is a strong evidence that aromatic amines require metabolic activation for carcinogenicity. The first step involves N-hydroxylation and N-acetylation, and the second step involves O-acylation yielding acyloxy amines. These compounds can degrade to form highly reactive nitrenium and carbonium ions. These electrophilic reactants may readily bind covalently to genetic material, namely cellular DNA and RNA (Brown & DeVito, 1993).

This process may induce mutations, and it is recognised that mutations can lead to formation of tumours.

Although the primary acute hazard associated with exposure to aromatic amines is carcinogenesis, methemoglobinemia is attributed to the same mechanism of metabolic activation.

Aromatic amines - structure activity relationship

For this class of organic compounds, the structure activity relationship between aromatic amines and carcinogenic potential has been reviewed in details (Milman & Weisburger, 1994).

Carcinogenic potential of aromatic amines varies considerably with the molecular structures, although the mechanism of metabolic activation, resulting in formation of electrophilic reactants, seems to be common. General trends are obvious and may outline a structure-activity relationship as follows:
Aromatic amines consisting of two or more conjugated or fused aromatic rings are associated with a high carcinogenicity potential. Single aromatic or non-conjugated ring amines may be carcinogenic too, but the potential is lower.
An aryl or alkyl group attached to the amino nitrogen can modify the carcinogenic potential by the interference of N-hydroxylation.
Certain substitution of the aryl ring has a fairly constant influence on carcinogenic potential. Aromatic rings substituted in para-position to the amino group are generally more carcinogenic than those non-substituted. Substitution with a methyl or a methoxy group in para- position to the aromatic amine group often enhances the carcinogenic potential, whereas sulphonic acid derivatives do not show mutagenic and carcinogenic potential.

Carcinogenic aromatic amines, which are common in industrial important azo dyes, are containing the moiety of:
aniline
toluene
benzidine
naphthalene

Correlation between exposure to aromatic amines containing the moieties mentioned above, and cancer in humans and/or in experimental animals has also lead to severe restriction or prohibition regarding manu- facture and use of these compounds.

Manufacture and use of azo dyes based on any of the 22 aromatic amines, presented in Table 5.3, have been restricted in several countries (Specht & Platzek, 1995). In Germany these amines are on a list encompassing hazardous substances in the working environment, see Table 5.3.

Table 5.3
Aromatic amines restricted according to the MAK- und BAT Werte Liste, 1998.
Regulerede aromatiske aminer i henhold til MAK- und BAT Werte Liste, 1998.

Moiety

Synonym

CAS

Aniline
4-chloroaniline 4-chloro-benzenamine

106-47-8

2,4,5-trimethylaniline 2,4,5-trimethyl-benzenamine

137-17-7

4,4´-methylenebis[ 2-chloro- aniline] -

101-14-4

4,4´-methylenedianiline 4,4´-diaminodiphenylmethane

101-77-9

4,4´-oxydianiline di (4-aminophenyl) ether

101-80-4

4-methoxy-m-phenylenediamine 4-methoxy-1,3-benzenediamine
2,4-diaminoanisol1

615-05-4

4-aminoazobenzene 4-(phenylazo)- benzeneamine

60-09-3

o-anisidine 2-methoxy-benzenamine

90-04-0

Toluene
o-toluidine 2-methyl-benzenamine

95-53-4

4-chloro-o-toluidine 4-chloro-2-methyl-benzenamine
4-chlor-o-toluidin1

95-69-2

5-nitro-o-toluidine 2-methyl-5-nitro-benzenamine
2-amino-4-nitrotoluidin1

99-55-8

6-methoxy-m-toluidine 2-methoxy-5-methyl-benzenamine
p-kresidin1

120-71-8

4-o-tolylazo-o-toluidine 4-amino-2´,3 dimethylazobenzene
o-aminoazotolulol1

97-56-3

4,4´-bi-toluidine 3,3´-dimethylbenzidine

119-93-7

4-methyl-m-phenylenediamine 4-methyl-1,3-benzendiamine
2,4-toluylendiamin1

95-80-7

4,4´-methylenedi-o-toluidine 3,3`-dimethyl-4,4´-diamino-
diphenylmethan1

838-88-0

Benzidine
benzidine 4,4´-diaminobiphenyl

92-87-5

4,4´-thiodianiline di(4-aminophenyl)sulphide

139-65-1

3,3´-dichlorobenzidine -

91-94-1

3,3´-dimethoxybenzidine o-dianisidine

119-90-4

biphenyl-4-ylamine 4-aminobiphenyl

92-67-1

Naphthalene
2-naphthylamine -

91-59-8

1 German synonym.

Problems of impurities

Several impurities may be found in almost all commercial available azo dyes. Impurities may be introduced during the manufacturing processes or during the storage.

Azo dyes, based on aromatic amines, may contain these amines as impurities introduced during the manufacturing process. For example, azo dyes based on benzidine or o-toluidine may contain residues of benzidine or o-toluidine, respectively, used as intermediates in the manufacturing process.

Aromatic amines may also be present as a result of thermal or photochemical degradation of azo dyes. It is known, that sunlight may cause release of 1-aminonaphthalene formed azo dyes based on this amine (Brown & DeVito, 1993).

Exposure

Exposure to azo dyes also entails exposure to the component aromatic amines due to:
breakdown of azo dyes.
presence of aromatic amines as impurities (their intermediates or breakdown products).

Exposure to aromatic amines is of great concern, as many of them are characterised by having serious long-term effects.

Exposure to azo dyes may take place through inhalation and accidental ingestion. Absorption of azo dyes through the skin is rather doubtful, whereas the aromatic amines may be absorbed.

In Denmark, occupational exposure to azo dyes may take place within colouring of textiles, leather and plastics.

Non-occupational exposure to azo dyes may take place by the wearing of coloured textiles and by playing with coloured toys which not conform to requirements and standards harmonised at the European level by the Council Directive concerning safety of toys.

Inhalation of cigarette smoke represents the greatest non-occupational exposure, as the smoke contain aromatic amines along with many other hazardous compounds. It is known that inhaled cigarette smoke enhance the incidence of bladder cancer, and heavy cigarette smoking doubles the risk of getting bladder cancer (Cartwright, 1983).

Summary

The acute toxicity of azo dyes is low. However, potential health effects are recognised, i.e. LD50 values between 250 and 2,000 mg/kg body weight.

Despite a very broad field of application and exposure, sensitising properties of azo dyes have been identified in relatively few reports. Red azoic dyes have been linked to allergic contact dermatitis in heavily exposed workers. Furthermore, textiles coloured with disperse azo dyes have caused allergic dermatitis in a few cases.

The azo linkage of the azo dyes may undergo metabolic cleavage which results in free component aromatic amines. After cleavage of the azo linkage, the component aromatic amines are absorbed in the intestine and excreted in the urine. 22 of the component amines are recognised as potential human carcinogens, and/or several of them have shown carcinogenic potential on experimental animals. Sulphonation of the dye reduces the toxicity by enhancement of the excretion.

Although the metabolic cleavage of azo dyes is the main source of aromatic amines, aromatic amines may also be present as impurities in commercial available azo dyes.

Due to a strong relationship between exposure to azo dyes and/or aromatic amines and evidence of human cancer, aromatic amines are the greatest hazard to health. Consequently, exposure to azo dyes based on aromatic amines, which are known or suspected human carcinogens, encompasses the greatest risk to health.

There is a small but possible risk of exposure to potential carcinogenic aromatic amines from dyes and coloured products in Denmark. Occupational exposure to azo dyes may take place in association with the colouring of textiles, leather and plastics. Non-occupational exposure may take place by wearing textiles, playing with toys and by inhalation of cigarette smoke. The exposure may take place as a result of a break- down of the dyes or due to impurities of the dyes.

Environmental fate and exposure

Releases into the environment

Dyes

Measured data concerning the emissions of azo dyes to the environment in Denmark are not available. This applies both for the production (processing) and the use phases.

The major route of release during the production phase is through waste water effluent from the processing industries, mainly from textile and to a smaller extent from leather. In the present survey it is assumed that releases from the remaining trades: paper mills, printing, plastics and paint industries are negligible and approximately 0. In addition, it shall be noted that no manufacture of dyes takes place in Denmark.

There is a potential release of dyes to the waste water during the consumption (use phase) of the end-products (paints, varnishes, textiles etc.) from industries as well as private households. However, the predominant potential release route from end-use is from waste deposited in landfills.

The potential atmospheric release route may be through particulate matter from soils which are treated with sludge, from waste deposits (land-fills), from incineration of waste and from emissions of the processing industry. It is estimated that the atmospheric release route is insignificant and approximately 0.

Agricultural soil fertilised with sludge may give rise to releases of dyes to soil/groundwater. In addition, landfill deposit of dyes contained in products may cause release of dyes to soil/groundwater, too.

The estimated Danish releases are shown in Table 5.4. The preconditions for the estimates are given in chapter 4. It should be noted, that the release to landfills is assumed to be associated, exclusively with the consumption of end-products (use phase).

Table 5.4
Estimated environmental releases of azo dyes in Denmark.
Estimeret frigivelse af azofarvestoffer til miljøet i Danmark.

Processing

Waste water (Influent, stp1)

Landfills

industry

Processing

Use

Use

 

tonnes

tonnes

tonnes

Leather

1

n.a.

9

Paint

n.a.

n.a.

n.a.

Paper

n.a.

n.a.

n.a.

Plastic

n.a.

n.a.

26

Print

n.a.

n.a.

n.a.

Textile

70

72

191

1 stp = Sewage Treatment Plant
n.a. = negligible amount.

Metabolites

Impurities of the dyes as well as decomposition by reductive cleavage of the azo dyes may result in transformation of the azo dyes to the degradation products/metabolites (aromatic amines), of which some are potentially carcinogenic. Estimation of the decomposition of azo dyes in the environment may be derived from knowledge of the structural and molecular composition of the azo dyes and of a stoichiometric equation.

The environmental exposure routes of the aromatic amines are essentially the same as the ones described for the dyes.

Degradation

Abiotic degradation

An important natural abiotic degradation mechanism is photolysis and hydrolysis as a function of pH in the range of pH 4-9 (ETAD, 1992a).

The evidence of the role of hydrolysis in degradation of azo dyes is not conclusive. Hydrolysis is by Baughman and Perenich (1988b) not considered to be important. If the dye is not broken during rigors of biological waste treatment, it is unlikely to degrade rapidly in the less severe conditions of the environment. This is supported by Clarke and Anliker (1980), who states that the reductive cleavage of the azo-bond is the major degradation pathway for azo dyes.

For the reactive dyes the abiotic half-life due to hydrolysis is approximately 2 days (IUCLID).

Photo-reduction of azo dyes to hydrazines and amines is possible, but it is likely to be very slow, except in oxygen-poor water. The stability of the dyes to visible and UV-light is very high, and therefore only slow degradation has been shown (Clarke & Anliker, 1980).

The photo-stability of azo dyestuffs is high in pure water but in the pre- sence of natural humic materials, the photo decomposition is strongly accelerated, probably through oxidation by single oxygen or oxy-radicals (Brown & Anliker, 1988).

Shu et al. (1994) demonstrated photo-oxidation (UV/H2O2- -photo chemical reactor) of two non-biodegradable azo dyes in waste water (Acid Red 1* and Acid Yellow 23*). It was observed that the decomposition of both azo dyes was pseudo-first order reactions with respect to azo dye concentrations. The reaction rates were dependent on the pH, the initial dye concentration and the hydrogen peroxide dosage, e.g. with high concentrations of H2O2 (18.95 mM) the half-life was 6 minutes for Acid Red 1* (20 ppm).

Other advanced oxidation processes include Fenton’s reagent and TiO2 photo-oxidation (Shu et al., 1994). A feasibility study by Dieckmann et al. (1994) indicates that azo dyes (Solvent red 1 and 4-hydroxyazoben-zene) can be degraded via sensitised photocatalysis on a surface of TiO2.

Shu and Huang (1995) investigated 8 acidic azo dyes for degradation of UV/Ozone. They found that the degradation rate were of the first order with respect to both azo dyes and ozone concentrations. UV-light did not significantly enhance the degradation ability. The half-lives were in the range of 1.2 to 2.6 minutes.

It is assumed that the main abiotic removal mechanism for dyestuffs in wastewater treatment plants is adsorption of sludge. However, other effects like sedimentation, precipitation or flocculation may also play a role (Pagga & Taeger, 1994).

Anionic dyes may be expected to react with Ca, Mg etc. to form highly insoluble salts (i.e. pigments) and thereby reduce the concentration, which is available for other reactions or biological effects (Baughman & Perenich, 1988b).

Other physico-chemical processes are flocculation, flotation, membrane filtration, electrokinetic coagulation, electrochemical destruction, ion-exchange, chemical oxidation and different sorption techniques. A review of the different treatment technologies and techniques and their efficiency towards degradation of xenobiotics has been given by Matsumoto et al. (1995). However as Banat et al. (1996) conclude, not one specific treatment process seems to be able to handle decolourisation of all textile waste waters. Generally, a customised process, which probably involves a combination of different methods, will be more applicable. Ozonation has achieved the greatest practical importance for removal of colours, but also precipitation and flocculation procedures have given good results. Decolourisation with reductive agents such as hydrosulphite is a workable proposition (Clarke & Anliker, 1980).

Metabolites

Some of the aromatic amines may be susceptible to photolysis, e.g.

4-methyl-m-phenylamine (HSDB, 1998).

Hydrolysis is, generally, not an important route of degradation of the aromatic amines (HSDB, 1998).

Summary

Even though the dyes have absorption maxima in the range of visible and UV-light, photo-reduction does not play a dominant role in the environmental fate of dyes, although its contribution to the total mineralisation of widely dispersed trace amounts may be underestimated. Furthermore, hydrolysis seems not to be an important degradation pathway either, except for reactive dyes, which are hydrolysed rapidly in aqueous solution.

For the metabolites, photolysis may be of some importance, whereas hydrolysis not seems to be an important degradation route.

Biodegradation

Razo-Flores et al. (1997a) estimate that due to the recalcitrance of azo dyes in aerobic environments, the azo dyes eventually end up in anaerobic sediments, shallow aquifers and in groundwater.

Extensive tests indicate that dyes are generally adsorbed to the extent of 40-80% by the biomass and are thus partly removed from the water phase in sewage treatment plants. They are, however, not biodegraded at this stage to any significant extent (Clarke & Anliker, 1980).

Dyes to be useful must possess a high degree of chemical and photolytic stability which implies that removal of dyes from effluents is difficult. Stability against microbial attack is also a required feature of azo dyes (Pagga & Brown, 1986). Subsequently, they are less amenable to biodegradation (Banat et al., 1996). It is thus unlikely that they, in general, will give positive results in short-term tests (e.g. OECD) for aerobic biodegradability (Brown & Anliker, 1988).

Furthermore, the electron-withdrawal character of azo-groups generates electron deficiency and thus makes the compounds less susceptible to oxidative catabolism. As a consequence, many of these chemicals tend to persist under aerobic environmental conditions (Knackmuss, 1996).

Biodegradation of azo dyes can occur in both aerobic and anaerobic environments. In both cases, the initial step in the biodegradation is the reductive cleavage of the azo-bond. Under aerobic conditions the initial step of cleavage of the azo-bond is typically followed by hydroxylation and ring opening of the aromatic intermediates (Zissi & Lyberatos, 1996).

Permeability through the cell wall has often been found to be the rate-limiting step in the reduction process. The microbial reduction of azo dyestuffs are either by reduction of living cells or by cellular extracts (Brown & Anliker, 1988).

Anaerobic and aerobic metabolic activities are a prerequisite for the complete biodegradation of recalcitrant aromatic pollutants, which contain electron-withdrawal substituents, such as azo dyes. Therefore, the recalcitrant nature of azo dyes can be overcome by utilising anaerobic-aerobic co-cultures (Field et al., 1995). This is supported by Clarke and Anliker (1980), who furthermore state that physical and chemical treatment is required as well. With the possible exception of basic dyes, the biological treatment processes (activated sludge) have in most cases proved to be insufficient for removal of dyestuffs from waste waters (Clarke & Anliker, 1980).

Bacteria - anaerobic

Brown and Laboureur (1983b) investigated the primary biodegradation of 13 azo dyes in an anaerobic sludge inoculum. The dyes were selected as commercially significant and represented both monoazo, diazo and polyazo dyes. The monoazo dyes were Mordant Blue 13, Mordant Black 9, Basic Red 18, Acid Yellow 151 and the diazos Direct Red 7, Acid Red 114*, Direct Blue 15*, Direct Yellow 12, Reactive Black 5 and Acid Blue 113*. All of these were substantially biodegraded (75-94%), whereas the polyazos Direct Black 19 and Direct Black 22 were only decolourised between 51-61% in a time period of 0 to 42 days.

Later results by Brown and Hamburger (1987) have confirmed that azo dyes are likely to undergo primary biodegradation in an anaerobic environment. The decolourisation was more than 90% in the time range of 0 to 56 days. The dyes tested were Acid Orange 7*, Acid Yellow 25*, Acid Yellow 36*, Acid Yellow 151, Acid Red 114*, Acid Black 24, Direct Red 7, Direct Blue 14*, Direct Blue 15*, Direct Yellow 12, Direct Yellow 50*, Mordant Black 9 and Mordant Black 11. This was also confirmed by Boethling et al. (1989) for Direct Red 28*.

Shaul et al. (1991) also found evidence of biodegradation of Acid Orange 7*, Acid Orange 8 and Acid Red 88. In 24 hours, 81 to 86% were degraded. The presence of sulfo groups on the aromatic component of some azo dyes seemed to inhibit the biodegradability significantly.

Direct dyes (Direct Red 28*, Direct Blue 1* and Direct Blue 14*) are degraded with more than 90% in anaerobic sediment-water systems with half-lives ranging from 2 to 16 days. The degradation is inhibited when the dyes are strongly bound to the sediment (Weber, 1991)

In sediments, Yen et al. (1991) showed that the degradation of two disperse azo dyes (Disperse Red 1 and Disperse Red 5) had half-lives within hours when the concentrations were kept below 10 ppm in the sediment. The reduction of nitro groups to amino groups and/or cleavage of the azo groups to give nitroanlines were found to be major pathways.

Zissi and Lyberatos (1996) demonstrated that Bacillus subtilis is, at least partly, able to degrade the disperse azo dye p-aminobenzene under anoxic conditions growing in a batch-reactor. The results proved that Bacillus subtilis co-metabolises p-aminobenzene under denitrifying conditions in the presence of glucose as a carbon source, producing aniline and p-phenyldiamine, as the N=N double bond is broken.

Other authors have reported degradation of disperse dyes with half-lives in order of minutes as well, e.g. Disperse Blue 79* (Weber & Adams, 1995; Freeman et al., 1996) and Disperse Red 1 with a half-life of less than 8 hours (Baughman & Weber, 1994).

The non-ionic dye Solvent Red 1 has been reported to have a half-life of 2.2 to 4 days (Baughman & Weber, 1994).

The reduction of benzidine azo dyes to free benzidine by soil bacteria has been reported for four aminobenzene azo dyes. The soil bacteria are Pseudomonas cepacia and Pseudomonas sp.. The initial reaction was azo reduction and cleavage, followed by acetylation and aromatic ring hydroxylation. The azo dyes were reduced with 42 to 91% at an aqueous concentration of 5 to 30 ppm during 24 hours of incubation. Similarly, a Plesiomonas bacterial species isolated from textile waste water has shown to degrade 5 different azo dyes under anaerobic conditions. Mixtures of sewage and soil bacteria (e.g. Pseudomonas aeuginosa) may also effectively degrade azo dyes. The dyes undergo azo-bond cleavage followed by carboxylation, hydroxylation and acetylation metabolism of the initial aromatic amine azo-reduction metabolites (Brown & DeVito, 1993).

Examples of removal of dyes in use in Denmark under anaerobic conditions are summarised in Table 5.5 below.

Table 5.5

Removal of azo dyes used in Denmark under anaerobic conditions.

Fjernelse af azofarver anvendt i Danmark under anaerobe forhold.

Chemical class

Period

Degree of removal

 

days

%

Acid Blue 113

0-421

94

Acid Orange 7

282

13

972

813

Acid Red 114

72

0-421

1002

621

Acid Yellow 25

562

572

Acid Yellow 36

72

972

Direct Black 19

0-421

511

Direct Blue 1

164

504

Direct Blue 14

72

34

>902

504

Direct Blue 15

0-421

831

Direct Red 28

44

504

Direct Yellow 50

352

1002

Solvent Yellow 1

135

895

Solvent Yellow 2

75

1005

1 Brown and Laboureur (1983b).
2 Brown and Hamburger (1987).
3 Shaul et al. (1991).
4 Weber (1991).
5 HSDB (1998).

Bacteria - aerobic

Like dyes in general, the hydrolysed dyes are practically not biodegraded in the short retention time of the aerobic treatment processes. Most dyes are degraded under anaerobic conditions. Such conditions are met in the anaerobic digestion process at sewage treatment plants, and in sediments and soils (ETAD, 1991).

Pagga & Brown (1986) tested 87 dyes in a short-term aerobic biodegradation based on the OECD Guideline for a static test method with activated sludge. They found no significant biodegradation, but substantial colour removal was observed which was attributed to the elimination of the dyes by adsorption. The tested dyes represented all the ionic characters and chemical types.

A study by Zhang et al. (1995) revealed that Acid Orange 7* and Acid Orange 8 can be degraded aerobically in a rotating drum biofilm reactor. The more complex Acid Orange 10* and Acid Red 14*, however, were not aerobically degraded. However, the authors demonstrated that cleavage of the azo bond occurred easily under anaerobic/anoxic biofilm conditions.

Knackmuss (1996) suggests that a total biodegradation of azo dyes may be accomplished by bacteria, harbouring a highly efficient uptake and an azo reductase system which are used in a two-step anaerobic/aerobic process, at least with regards to biodegradation of sulphonated naphthalenes.

Fungi

Microbial degradation of lignin-containing pulp and paper waste water has been demonstrated by several authors, especially with the white-rot Basidiomycete fungus: Phanerochaete chrysoporium. The mechanism of colour removal involves lignin peroxidase and Mn-dependent peroxidase or laccase enzymes. The degradation of azo dyes is apparently dependent on the availability of nitrogen. If there is a high concentration of N the degradation rate decreases. Banat et al. (1996) have reviewed the literature and found out that azo dyes may be degraded by the fungus between 23 and 90% in a time span of 3 to 21 days with different concentrations. A wide variety of dyes has been tested, among them Acid Red 114*, Acid Red 88, Direct Blue 15*, Disperse Yellow 3, Disperse Orange 3 and Solvent Yellow 14* (Spadaro et al., 1992). In addition, other anionic dyes, like Reactive Orange 96, Reactive Yellow 5 and Reactive Black 5, have been demonstrated to be biodegraded by the white-rot fungus Phanerochaete chrysosporium, too (Heinfling et al., 1997).

The actinomycete strains, mainly streptomycetes, isolated from soil samples have been demonstrated to decolourise effluents containing different types of reactive dyes. In a study carried out by Zhou and Zimmermann (1993) it was concluded that the decolourisation of Reactive Red 147 was due to adsorption rather than biodegradation. Banat et al. (1996) has reviewed the studies of other fungal biodegradation of azo dyes and several (7 in total) other species have shown to decolourise but mainly by way of adsorption.

There are conflicting evidence of the influence of the substituents on the aromatic ring with regards to the effect on biodegradability by Phanerochaete chrysoporium. Paszczynski et al. (1992) found that substitution with sulfo groups on the aromatic component of some azo dyes not seemed to affect the biodegradability of the anionic azo dyes significantly. Pasti-Grigsby et al. (1992), however, found that significant degradation of the azobenzene derivative dyes and naphtol-derivative dyes (e.g. Acid yellow 9 and Acid Orange 12 (anionic)) occurred solely when the hydroxy group was in a specific position relative to the azo linkage. Spadaro et al. (1992) showed that when the aromatic rings of the neutral dyes (Solvent Yellow 14*, Disperse Orange 3 and Disperse Yellow 3) had substituted hydroxyl, amino, acetamido or nitro groups, the mineralisation was greater than by those with unsubstituted rings.

Algae

Decolourising with algal cultures has been found by Jinqi and Houtian (1992). The reduction of algae resembles that of the bacteria. The azo reductase of the algae Chlorella and Oscillatoria is responsible for degrading azo dyes into aromatic amines. The aromatic amine is then subject to further degradation by the algae. As for bacteria, azo compounds with a hydroxy or an amino group are most likely to be readily degraded than those with a methyl, methoxy, sulfo or a nitro group.

Mineralisation

FitzGerald and Bishop (1995) found an almost total decolourisation in the first stage of an anaerobic/aerobic treatment of sulphonated azo dyes (Acid Orange 10*, Acid Red 14* and Acid Red 189). Analyses of the intermediates at the first and second stages (aerobic) showed virtually no concentration of intermediates, which may indicate a total anaerobic mineralisation. In contrast, Seshadri et al. (1994) found that the aromatic amines remained undegraded in an anaerobic fluidised bed reactor.

Razo-Flores et al. (1997b) have demonstrated that Mordant Orange 1 may be completely degraded (mineralised) in a continuous upward-flow anaerobic sludge bed reactor in the presence of co-substrates.

Razo-Flores et al. (1997a) have further demonstrated that the azo dye, azodisalicylate is completely biodegradable in the absence of oxygen. The dye is mineralised in an adapted methanogenic consortium to CH4 and NH3 in both batch assays and continuous bioreactors.

Degradation of metabolites

Free aromatic amines are generally susceptible to environmental degradation (Brown & DeVito, 1993). Zerbinati et al. (1997) have found that naphthalenesulfonates can undergo oxidative degradation under physico-chemical conditions similar to those occurring in a river. However, other studies have shown that, e.g. benzidine is bound with the humic acid fraction of the soil (Weber, 1991).

Brown and Laboureur (1983a) showed in aerobic biodegradation tests that the four aromatic amines: aniline, p-anisidine, p-phenetidine and

o-toluidine are ready biodegradable and that both o-anisidine and

3,3´-dichlorobenzidine are inherent biodegradable in accordance with the OECD test guidelines. Brown and Hamburger (1987) confirmed these results for the lipophilic aromatic primary amines, but depending on their precise structure, some sulphonated aromatic amines may not be degradable.

Under aerobic conditions another type of recalcitrance can be recognised, namely, the tendency of certain compounds, susceptible to free radical reactions, to undergo oxidative coupling. These coupling reactions can result in the formation of recalcitrant humic-like polymers or in irreversible covalent binding of the pollutant into the soil humus. Aromatic amines and nitroaromatics are susceptible to these polymerisation reactions. Formation of azo compounds by oxidative coupling has been demonstrated in aerobic enrichment cultures from the aromatic amines (Field et al., 1995)

The metabolites of aromatic primary amines are not rapidly degraded under anaerobic conditions (Brown & Hamburger, 1987). Electron donating amino groups are expected to pose a serious problem to further reductive biotransformations by anaerobes. However, there is evidence for anaerobic aniline biodegradation by sulphate reducing bacteria and in mixed cultures under denitrifying conditions. Aniline degradation by a methanogenic consortium has also been claimed. Aromatic amines with carboxy, hydroxy and methoxy substituents are potentially mineralisable under methanogenic conditions (Field et al., 1995). Another example is o-toluidine, which is not degraded under anaerobic conditions (HSDB, 1998).

Summary

Various microbial species, i.e. fungi, bacteria and algae may be able to biodegrade azo dyes in an anaerobic environment. Total mineralisation or further degradation of the metabolites may predominantly take place in an aerobic environment.

The universal degradation route seems to be initial reductive cleavage of the azo bond followed by e.g. acetylation, carboxylation and aromatic ring hydrolysation.

The rate limiting step for bacterial degradation is the uptake across cell membranes for intracellular reduction, whereas some fungi may degrade the dyes extracellularly.

The substituents and the substitutional pattern may also significantly influence the biodegradability. The reported effects are contradicting, but the ionic azo dyes with hydroxy or amino groups are most likely to be readily biodegraded, compared to those with methyl, methoxy, sulfo or nitro groups. For the non-ionic dyes (disperse, solvent and mordant) an enhanced biodegradation is observed with hydroxyl, amino, acetamido or nitro groups compared to unsubstituted rings.

It is difficult to generalise about degradation rates and the degree of removal for specific azo dyes or for the different chemical classes based on the findings in the literature, because the experimental conditions vary.

However, biodegradation of azo dyes varies, in general, from hours to several months or more depending on, among other things, the physico-chemical properties of the dyes. The molecular size of the azo dyes, especially solvent and disperse dyes, may reduce the rate and probability of biodegradation. This is due to limited uptake possibilities, and the substituents may also influence the degradation rate.

The metabolites are primarily biodegraded under aerobic conditions. Some of the metabolites are ready biodegradable, and some of the sulphonated aromatic amines may not be degradable.

Distribution

Volatilisation

Data concerning the volatilisation of azo dyes from aqueous surfaces are not available. With respect to volatilisation it is prerequisite to distinguish between the ionic dyes and non-ionic dyes, because ionic compounds are generally non-volatile. (Brown & Hamburger, 1987). Therefore, volatilisation will not be important for acid, direct, basic and reactive dyes. In principle, the solvent and disperse dyes have the potential to be volatile, but as they are large, complex molecules they can be expected to have low vapour pressures. Another reason for volatilisation to be unlikely for the uncharged dyes is that the escaping tendency or fugacity, which drives volatilisation, is also the driving force for both sorption and bioconcentration (Baughman & Perenich, 1988b).

Baughman and Perenich (1988a) calculated Henry’s law constants from solubility and vapour pressure. The values show that the disperse dyes will be entirely vapour-phase controlled in the environment in their rate of volatilisation from water and that this process is extremely slow. The vapour pressures lie in the range of 2 ´ 10-14 to 1 ´ 10-6 mmHg and the solubility in the range of 2 ´ 10-9 to 4.5 ´ 10-6 mol/l. The Henry law constant is on average 10-10 atm ´ m3/mol for disperse dyes.

Metabolites

In general, the metabolites, i.e. the 22 potentially carcinogenic aromatic amines, show moderate to low volatilisation with Henry’s law constants in the range of 4.7 ´ 10-11 (3,3´-dimethoxybenzidine) to 2.0 ´ 10-6 atm ´ m3/mol (o-toluidine).

Summary

Due to the chemical characteristics of the azo dyes, volatilisation from surfaces of either water or soil (wet or dry) is considered to be insignificant for both ionic and neutral (non-ionic dyes). This applies for the metabolites, too.

Adsorption

Because of dyestuffs inherent high affinity to substrates, they are adsorbed onto the sludge during sewage treatment and are thus removed from the final treated effluent (Anliker, 1986). But due to the chemical composition of some of the dyes, they may pass the sewage treatment unaffected and thus end up in the aquatic environment. Extensive testing indicates that dyestuffs are generally adsorbed to the extent of 40-80% by the biomass and are thus partially removed in sewage treatment plants (Clarke & Anliker, 1980). However, due to their relatively low affinity to substrates, the removal of the hydrolysed dyes (e.g. Reactive dyes) by adsorption onto the sewage sludge is only in the range of 0-30% (ETAD, 1991).

In the practical concentration range of 10 to 50 mg dye/l, there is an almost linear relationship between the concentration in solution and the amount adsorbed. The adsorptive capacity of activated sludge for dyes investigated was, in neutral media, in the range 0.01 to 4% of dyestuff on dry weight sludge (Clarke & Anliker, 1980).

The chemical properties and substitutional pattern of the chemical structure of the dyes and the composition of the waste water influences the degree of adsorption. The adsorption depends on the pH, salinity and the concentration and nature of organic contents.

Based on the properties of sediments, cation exchange is anticipated to be extensive and rapid for the basic dyes. A similar situation should exist for the anionic acid and direct dye, but the equilibrium constants would probably be much smaller (Baughman & Perenich, 1988b).

Shaul et al. (1991) investigated the partitioning of water-soluble azo dyes in the activated sludge process. A total of 18 dyes were tested and categorised according to their behaviour in the tests (Table 5.6). For Group 1 it was concluded that the high degree of sulphonation enhanced their water solubility and limited their ability to adsorb onto the biomass. Although the dyes in Group 2 were highly sulphonated, their greater molecular size was thought to account for their greater degree of adsorption.

Table 5.6

Fate of water-soluble dyes in activated sludge.

Vandopløselige farvestoffers skæbne i aktivt slam.

Group 1: Group 2: Group 3:
Dyes passing through essentially unaffected Dyes removed by adsorption Dyes showing evidence of biodegradation
Acid Black 1* Acid Blue 113* Acid Orange 7
Acid Orange 10* Acid Red 151 Acid Orange 8
Acid Red 1* Direct Violet 9 Acid Red 88
Acid Red 14* Direct Yellow 28  
Acid Red 18    
Acid Red 337    
Acid Yellow 17*    
Acid Yellow 23*    
Acid Yellow 49    
Acid Yellow 151    
Direct Yellow 4    
Ref.: Shaul et al. (1991).

Weber (1991) has demonstrated that the sorption of several weakly basic benzidine-based dyes (Direct Red 28* (disulphonated) and Direct Blue 14* (tetrasulfonated)) strongly depend on the pH and the nature and concentration of inorganic salt in solution in an anaerobic sediment-water system. Sorption is strongly favoured with decreasing pH and increasing salt concentration. The sorption was enhanced especially for Direct Red 28*, which was less substituted.

Pagga and Taeger (1994) have found that the colour elimination of acid and disperse azo dyes (Acid Orange 7*, Acid red 88, Disperse Orange 29 and Disperse Yellow 5) depends on the hardness of the water. A high concentration of calcium ions favours adsorption as well as flocculation or precipitation processes or a better settling of the sludge and less turbidity.

In a study by Yen et al. (1990), it has been shown that newer disperse dyes show a higher degree of partitioning into the sediment than older disperse dyes based on calculated sediment concentrations.

Metabolites

The metabolites adsorb, except for 4-methyl-m-phenylenediamine, mode- rately to strongly onto sediments and soil. 4-methyl-m-phenylenediamine

does not adsorb to any significant degree (HSDB, 1998).

Summary

The removal of various dyes from different classes has been studied and the removal pattern may be summarised as shown in Table 5.7.

Table 5.7

Removal patterns of various classes of dyes.

Fjernelsesmønster for forskellige farvestoftyper.

Classes

Removal pattern

Acid High solubility leads to low adsorption, which appears to depend on the degree of sulphonation.
Basic Typically high levels of adsorption.
Direct High degree of adsorption, apparently unrelated to the number of sulphonic acid groups.
Disperse Adsorption in the high-to-medium range.
Reactive Very low degree of adsorption, apparently unaffected by the degree of sulphonation or ease of hydrolysis.

Extensive adsorption onto soil and sediment has been demonstrated in several experiments. It is concluded that adsorption is the major route of removal of dyes in the environment. Adsorption is an important removal pathway for the metabolites, as well.

A high degree of solubility and sulphonation reduces adsorption, whereas increasing molecular size, hardness of the water and salinity favour sorption. This applies for a decreasing pH, as well.

Bioaccumulation

The obtained data on bioaccumulation are primarily derived from fish-

tests.

The uptake rates are influenced by the partition coefficient (log Kow) (Erickson & McKim, 1990). Other factors may be of primary importance for the uptake as well, e.g. diffusional resistance, molecular size, respiratory volume and gill perfusion (Niimi et al., 1989).

The elimination rates for hydrophobic chemicals are low. For hydrophobic chemicals it has often been shown that uptake and clearance between fish and water is a first-order exchange process (Van Hoogen & Opperhuizen, 1988).

Anliker et al. (1981) have presented, estimated and experimentally assessed the log Kow and have experimentally assessed the log BCF (bio- concentration factor) in fish (MITI standard) for 50 azo dyes, representing both the ionic forms and the neutral forms. The average values for the different dyes are presented in Table 5.8 below.

Table 5.8

Partition coefficients and the measured bioaccumulation factors for 50 azo dyes allocated on 5 chemical (technical) types.

Fordelingskoefficienter og den modsvarende målte bioakkumulationsfaktor for 50 azofarvestoffer fordelt på 5 kemiske (tekniske) grupper.

Structural type

Partition coefficient

(log Kow)

Bioaccumulation Factor

( in fish)

(log BCF)

Acid

-3.3-0.01

0-0.7

Basic (only one)

-1.0

-0.3

Direct (only one)

<<0

0.2

Disperse

2.2-5.5

0-1.76

Reactive

-2.2- -0.4

-0.2-0.7

Ref.: Anliker et al. (1981).

The survey shows, with a few exceptions, that the very hydrophilic (ionic) dyes have a log BCF of - 1 to 1, although from the log Kow lower log BCFs may have been predicted. This is explained by the adherence of dyes to the outside of the fish or to the intestine. None of the dyestuffs bearing at least one charged group has showed a log BCF larger than 1. It has been demonstrated that disperse dyes do not bioaccumulate in fish even though their log Kow values were larger than 3. The molecular weight was relatively high, between 450 to 550 g/mol, making the transport across membranes difficult.

These findings have been confirmed in other studies. The partition coefficients of 21 reactive dyes were very low (log Kow < 0) and none of these dyes have showed any tendency of bioaccumulation in the flow-through tests (MITI-standard) in the carp. (ETAD, 1991). In Carrassius sp., which was exposed to 2 mg/l and 0.2 mg/l for 42 days, the BCFs were less than 1.1 and less than 11, respectively (IUCLID).

ICI has carried out eight-weeks accumulation studies on the Carp (Cypri- nus carpio) (MITI standard). The results indicate that neither the 30 water-soluble2 nor the 12 disperse dyes2 with exposure levels up to 10 mg/l were accumulated. For the soluble dyes the accumulation factor was below the detection limit. The low accumulation of the soluble dyes may be expected, and the low accumulation factors found for the disperse dyes may be due to their relatively high molecular weight (typically 300-500) or because their absolute fat solubility is relatively low (Brown, 1987).

Anliker and Moser (1987) have investigated the melting point, the log Kow, solubility in water and n-octanol and the log BCF in fish for 8 disperse dyes (nitroazobenzene and phenylazopyridone types):

molar weight:
360 to 546 g/mol
melting point:
117 to 225 0C
solubility in water:
n.d.
solubility in n-octanol:
81 to 2,430 mg/l
log Kow:
2.5 to 5.4 (majority above 3)
log BCF (exp.):
0.3 to 1.76

They found that the high Kow suggested strong bioaccumulation tendencies, but the bioaccumulation was below 100. It was hypothesised that this behaviour may be due to their pronounced aggregation tendency, making transport across membranes difficult. The findings of Opperhuizen et al. (1985) support this. Their results indicated that for extremely hydrophobic chemicals with an effective cross section over 9.5 Å, a lack of uptake into biota (fish) can be expected, as membrane permeation seems practically impossible.

A study of 75 disperse dyes, even highly lipophilic ones, by Anliker (1986) and a later study by Anliker et al. (1988) on 23 disperse dyes3, including highly lipophilic ones, confirmed the above mentioned observations.

Similar results have been reported for the BCFs of chlorinated aromatic amines in guppies (Poecilia reticulata). The experimentally demonstrated values of BCF are significantly smaller that the calculated values of BCF (Wolf et al., 1992).

The azo compound (not dye) 3,3´,4,4´-tetrachloroazobenzene (TCAB), a common contaminant from 3,4 dichloroaniline based herbicides and of agricultural soils, has been tested (short-term) on the aquatic snail Indoplanorbis exustus by Allison and Morita (1995a). They found that even at detrital exposures of 2,500 ppm, the maximum level only reached 287 ppb (whole body basis). The authors above (1995b) also came to the same result in the Japanese Medaka (Oryzias latipes). The fish inhabit still waters and paddy fields and their drains. The study showed that TCAB is bioadsorbed and to some extent bioaccumulated in the fish. The contaminant was administered through the food.

Apart from the above stated findings, only a small amount of data was found in the literature and databases on the log Kow and the log BCF for specific azo dyes included in the present survey. The results of the literature study are presented in Table 5.9.

Table 5.9

Partition coefficient and bioconcentration factor for azo dyes used in Denmark.

Fordelingskoefficient og biokoncentrationsfaktor for azofarvestoffer anvendt i Danmark.

Structural type

Partition coefficient

(log Kow,, est.)

Bioconcentration

factor ( in fish)

(log BCF)

Reference

Acid Orange 10

-4.6

0 (est.)

HSDB

Acid Red 114

-

1.6-1.9

MITI

Acid Yellow 23

-

-0.54-0.48

MITI

Direct Black 38

2.0

1.3 (est.)

HSDB

Direct Blue 1

-

0.3 (est.)

HSDB

Disperse Blue 79

4.79

4.09 (est.)

Yen et al., 1990

Solvent Red 24  

-0.54-1.04

MITI

Solvent Yellow 1

2.62

1.76 (est.)

HSDB

Solvent Yellow 2

4.05

3.25 (est.)

HSDB

Solvent Yellow 3

3.92

2.29-2.75 (est.)

HSDB

Compared to the findings of Anliker et al. (1981) shown in Table 5.8, the log BCF for Acid Red 114* is above the range reported, whereas Acid Yellow 23* is in agreement. The remaining dyes are incomparable, as they are based on estimates rather than actual experimentally measured values of BCF. The solvent dyes were not included in the Anliker et al. (1981) study.

However, the calculated log BCF value for Disperse Blue 79* is in agreement with the findings of Anliker et al. (1981), whereas Direct Black 38* is twofold higher than reported by the above authors.

Metabolites

Both measured and estimated log BCFs for the cleavage products of the dyes, i.e. the 22 potentially carcinogenic aromatic amines, are according to ECDIN and HSDB (1998) below 3. The highest estimated log BCF has been found for 4-o-tolylazo-o-toluidine (2.75) (HSDB, 1998). The majority lies in the range of 1.5 to 2.0. The lowest values (<1.47) have been reported for o-anisidine (0.85), o-toluidine (1.2) and 4,4´-methylenedi-aniline (1.1). The log BCF values for the aromatic amines indicate that there is a risk of biomagnification for a great majority of the metabolites.

Summary

For the compounds with log BCFs larger than 3, there is a high risk of bioaccumulation, whereas for compounds with log BCFs between 1.47 and 3, the risks of biomagnification and secondary poisoning are important. For compounds with log BCF values below 1.47, there is no immediate concern with regard to bioaccumulation (Franke et al., 1994).

When looking at the values of the dyes included in the present survey, it is indicated that Acid Red 114 may bioaccumulate in fish, whereas the remaining ionic dyes do not seem to have any significant bioaccumulation potential. However, the estimated log BCFs for the non-ionic dyes, i.e. disperse and solvent, indicate a potential risk of bioaccumulation.

The estimated values for log BCF are generally to high which several authors have found. Therefore, the evidence of the risk bioaccumulation of the non-ionic azo dyes must be further validated, taking the potential barrier of uptake into account, as a result of the high molecular size of these compounds.

Generally, the cleavage products of the azo dyes, i.e. the aromatic amines, have a potential for bioaccumulation, too.

Aquatic compartment

Monitoring data

Only a few monitoring studies of environmental levels of dyes have been found, and data from Denmark have not been obtained.

In a study conducted by the US EPA, effluents from 25 textile industries were measured. The average TOC was measured to 276 mg/l (range 55 to 1,120 mg/l). The dyestuff itself has not been measured, but it is estimated that the dye contributes between 2 and 10% of the TOC and COD indicating worst case levels of 5.5 to 112 mg/l. However, the typical dye concentration lies in the range of 10 to 50 mg/l. Decolourised effluents contain less that 1 mg/l dye, and the TOC contribution of dyestuff following the primary and biological treatment stages is normally considerably less than 0.5 mg/l. In the same study, effluents from a tannery were measured, and the raw effluent contained 22 to 56 ppm dyestuff (Clarke & Anliker, 1980).

The following Table 5.10 and Table 5.11 summarise additional data regarding environmental monitoring of dyes in water and sediment.

Table 5.10

Monitoring data of dye concentrations in water.

Moniteringsdata for farvestofskoncentrationer i vand.

Compound

Concentration

Location/Reference
 

Treated sewage effluent (ppb)

River

(ppb)

Reservoir for drin- king water (ppb)

 
Acid Blue 1a

12.3

1.7

0.6

Thames, Lee, UK

(Brown & Anliker, 1988)

Fluoroscent whiteningb  

0.8-8

n.d (detection limit 0.01).

European river (Anliker, 1986)
Acid Yellow 219b

120

22

5

Coosa River, US

(Brown & Anliker, 1988)

Acid Orangeb  

10

  Coosa River, US

(ETAD, 1992b)

Acid Redb  

2

  Coosa River, US

(ETAD, 1992b)

Disperse Blue 26b,c

1985

1986

 

9.95

3.8

  Yamaska River, Canada

(ETAD, 1992b)

Disperse Red 60b,c

1985

1986

 

3.3

n.d.

  Yamaska River, Canada

(ETAD, 1992b)

Disperse Blue 79*c

1985

1986

 

17.1

3.1

  Yamaska River, Canada

(ETAD, 1992b)

Solvent Yellow 1*

522,7

    Organics and Plastic industry, US (HSDB)
Most common acid dyesb

20

    Coosa River , US

(Anliker, 1986)

a: Not an azo dye compound.

b: The chemical class of the dye was not stated.

c: Improved waste water treatment was installed between the two years.

Table 5.11

Monitoring data of dye concentrations in sediment.

Moniteringsdata for farvestofskoncentrationer i sediment.

Compound

Sediment

Suspended

solids

Location/

reference

 

mg/kg dw

mg/kg dw

 
Dyesb

0.1-3

  Coosa River, US (Brown & Anliker, 1988)
No individual synthetic dyesb

n.d.

(detection

limit 0.05)

  Rhine,

(Brown & Anliker, 1988)

Disperse Blue 26c

1985

1986

1.1

2.9

4.6

6.7

Yamaska River, Canada

(ETAD, 1992b)

Disperse Blue 79*c

1985

1986

1.5

4.2

0.8

3.3

Yamaska River, Canada

(ETAD, 1992b)

a: Not an azo dye compound.

b: The chemical class of the dye was not stated.

c: Improved waste water treatment was installed between the two years.

Estimation of PEC for the aquatic compartment

In most industrialised countries only about 20% or less of the release from processes will reach open water due to effective adsorption in the primary and the biologic treatment stages (Clarke & Anliker, 1980).

However, in the present calculation of PEC effluent, stp, two scenarios will be presented.

The estimation of PEC effluent, stp is based on the following assumptions:
The processing industries do not treat waste water in agreement with TGD (1996).
Between 40 and 80% of the azo dyes are adsorbed in the sewage treatment plant (STP) (Clarke & Anliker, 1980). Resulting in a worst case scenario of an adsorption of 40% (60% release to the effluent) and a best case scenario of an adsorption of 80% (20% release to the effluent).
Adsorption is the only removal route of azo dyes in the STP, i.e. there is no abiotic or biotic degradation.

Furthermore, a standard STP scenario, in compliance with TGD (1996), is used. According to this standard the values presented in Table 5.12 are standard characteristics of a STP:

Table 5.12

Standard characteristics of a sewage treatment plant.

Standardkarakteristika for et rensningsanlæg.

Parameter Symbol Unit Value
Capacity of local STP Capacity stp [eq] 10,000
Amount of wastewater per inhabitant Waste inhab [lxd-1xeq-1] 200
Surplus sludge per inhabitant SURPLUSsludge [kgxd-1xeq-1] 0.011
Concentration susp. matter in influent SUSPCONCinf [kgxm-3] 0.45

Ref.: TGD (1996).

The calculation of PEC influent, stp is simplified and based on the equation below:

PEC influent, stp = Release wastewater /Waste inhab ´ Capacity stp ´ 365

The calculation of PEC effluent, stp is simplified and based on the assumptions mentioned above. In addition, the PEC effluent, stp for the processing industry is corrected for the number of sites present in Denmark , i.e. 40 sites for textile colouring and 1 site for leather dyeing. For the use, the number of inhabitants in Denmark (approximately 5 millions) is normalised to the capacitystp.

PEC effluent, stp = PEC influent, stp ´ (1- adsorption factor/ (number of sites) or inhabitants in Denmark (Table 5.13).

PEC surface water = PEC effluent, stp ´ dilution factor

According to the TGD (1996), the dilution factor is 10.

Table 5.13

Estimated PECeffluent, stp and PECsurface water for azo dyes.

Estimeret PECudløb, stp og PECoverfladevand for azofarvestoffer.

 

Re-
lease

PEC influent, stp

PEC
effluent, stp

PEC
effluent, stp

PEC
surface water

PEC
surface water

 

t/year

mg/l

mg/l/site or inhab.

mg/l/site or inhab.

mg/l/site or inhab.

mg/l/site or inhab.

Processing    

Worst case

Best case

Worst case

Best case

Textile

70

95.89

1.44

0.48

0.14

0.048

Leather

1

1.37

0.82

0.27

0.08

0.027

Use            
Textile

72

98.63

0.12

0.04

0.012

0.004

Leather

-

-

-

-

-

-

Total

143

-

-

-

-

-

The PECsediment is calculated from:

PECsediment = PECsurface water  adsorption factor

In Table 5.14, the PECsediment is presented.

Table 5.14

Estimated PECsediment for azo dyes.

Estimeret PECsediment for azofarvestoffer.

Scenario

PEC surfacewater

Adsorption factor

PECsediment

 

mg/l

 

mg/kg

Worst case      
Processing      
Textile

0.14

0.8

0.11

Leather

0.08

0.8

0.06

Use      
Textile

0.01

0.8

0.008

Leather

-

0.8

-

Best case      
Processing      
Textile

0.05

0.4

0.02

Leather

0.03

0.4

0.01

Use      
Textile

0.004

0.4

0.002

Leather

-

0.4

-

Concerning the concentration of azo dyes in the sludge, the estimation is based on that the production of sludge amounts to 170,000 tonnes dw/year in Denmark (Miljøstyrelsen, 1996b). The "worst case" of adsorbed azo dyes onto the sludge is 80% and the "best case" is 40%. The calculated concentration in sludge is based on the following equation:

PEC sludge = (Release ´ adsorption factor ´ 106/ Sludge rate)/(number of sites) or inhabitants.

Sludge rate = 170.000 tonnes/year

Table 5.15

Estimated PEC sludge for azo dyes

Estimeret PECslam for azofarvestoffer.

 

Release

PECsludge

PECsludge

 

t/year

mg/kg/site or inhab.

mg/kg/site or inhab.

Processing  

Worst case

Best case

Textile

70

8.24

4.11

Leather

1

4.70

2.45

Use      
Textile

72

0.68

0.34

Leather

-

-

-

Total

143

-

-

The estimated PECsurface water for processing and use is in the range of 0.04 to 1.44 mg/l. According to the monitoring studies (Table 5.10) a range of 0.012 to 0.523 mg/l for treated sewage effluent has been found. If comparing the two, the estimated PECeffluent, stp is approximately 3 times higher. Compared to the concentrations found in river water, the estimated PECsurface water is 2 to 6 times higher which may be due to the dilution effect. The estimated PECsediment is, on the other hand, below the range of the monitored data (Table 5.11), namely 0.002 to 0.11 mg/kg dw.

Due to the lack of monitoring data of environmental concentrations of azo dyes in Denmark, it is not possible to validate the estimated PECs based on Danish data. The basic assumption, however, that the processing industries do not carry out waste water treatment prior to outlet (PECinfluent, stp) is unlikely, because most of these companies, if not all, are encompassed by a special section of the Danish Environmental Protection Law (chapter 5). Hence, their emissions are restricted and must be approved by the authorities. Subsequently, the companies are obliged to have some degree of waste water treatment prior to the outlet to the municipal STP.

Assuming that 40 to 80% of the dyes are removed from waste water before the outlet from the industry and likewise in the STP, this indicates that the actual PECeffluent, stp for the processing and use phase is more likely to be in the range of 0.024 to 0.864 mg/l and the PECsurface water in the range of 0.002 to 0.086 mg/l. These concentrations are within the same range, and for PECsurface water approximately 4 times higher compared to the findings in the aforementioned monitoring studies. This indicates at least for the best case scenario, that the estimated PECs may be realistic.

If it is estimated that the PECsurface water is too high, then the PECsediment has to be reduced in the same order of magnitude. Resulting in a concentration of 0.001 to 0.090 mg/kg from processing and use which is within the low range of the monitoring studies (Table 5.11).

If it is assumed that the companies carry out waste water treatment, the PECsludge, stp may also be reduced 2 to 5 times, depending on the degree of adsorption (40-80%) at the companies, and this results in a range of 1.18 to 5.62 mg/kg for processing and use.

Atmosphere

Monitoring data

No data have been obtained concerning monitoring of azo dyes in the atmosphere or bound to particulate matter.

Estimation of PEC for the atmosphere

It has not been attempted to calculate the atmospheric PEC, but it is estimated that the PEC is very low, because volatilisation is highly unlikely for the azo dyes from both moist and dry surfaces. Furthermore, the release from the processing industry and from incineration is considered to be very low (approximately equal to 0).

Terrestrial compartment

Monitoring data

There are no direct route by which agricultural soils may become contaminated with synthetic dyes. In principle, it is possible that the disposal of sludge from sewage treatment plants, which receive dye-house effluents, may provide an indirect route of exposure. Although, there appear to be no reported data of the levels of dyestuffs on agricultural soils, estimates based on the principles elaborated by the OECD, would indicate a worst case level of 1 mg/kg (w/w of dry soil) (Brown & Anliker, 1988).

Furthermore, deposition of particulate matter may be a potential pathway for the terrestrial environment, but as stated above it is considered to be an unlikely pathway.

A practical demonstration has showed that sewage sludge contaminated with dyes, when held under simulated landfill conditions, does not release dyes into the leachate. The amine metabolites, which may be expected to be produced from these dyes, cannot be found in the leachate or interstitial water either (Brown & Anliker, 1988)

However, no data have been obtained on terrestrial monitoring of azo dyes.

Estimation of PEC for the terrestrial compartment

The sources of environmental releases of azo dyes in the terrestrial environment are waste disposal in landfills and sludge applied to agricultural soil.

It is estimated that the total amount of sludge per year in Denmark is 170,000 tonnes of dry weight. About 114,000 tonnes (67%) are used in agriculture and 20,000 tonnes (12%) are deposited in landfills. The rest (21%) is incinerated (Miljøstyrelsen, 1996b).

It is not known how many hectares of agricultural soil which are fertilised with sludge in Denmark. But according to the TGD (1996), the following characteristics of soil and soil uses are accepted:

Table 5.16

Standard environmental characteristics for soil.

Standard miljøkarakteristika for jord.

 

Depth of soil

Rate of sludge application

 

[m]

[kgdwtxm-2xyear-1]

PEC localagr.soil

0.20

0.5

Ref.: TGD (1996).

In section 2.3.4 of the TGD (1996), the standard environmental characte- ristics are defined, and on this basis it may be calculated that the density of the soil is 1.7 t/m 3. By application of the depth of soil of 0.2 m in accordance with the TGD (1996), it is estimated that the weight of soil per square meter is equal to 0.34 tonnes.

Subsequently, assuming that in a worst case scenario 80% of the azo dyes are adsorbed onto the sludge, and that in a best case scenario 40% are adsorbed onto the sludge, then the amount of azo dyes on the agricultural fields can be estimated from the following equation:

PECagri sludge = (release ´ adsorption factor ´ fraction to agriculture)/ (sludge amount/application rate) ´ soil weight.

Table 5.17

Estimated PECagri sludge for azo dyes.

Estimeret PECagri slam for azofarvestoffer.

 

Release

PECagri sludge

PECagri sludge

 

tonnes/year

mg/kg

mg/kg

Processing  

Worst case

Best case

Textile

70

0.484

0.242

Leather

1

0.007

0.003

Use      
Textile

72

0.498

0.249

Leather

-

-

-

Total

143

-

-

The allocation of sludge to landfill disposal amounts to 20,000 tonnes (dw)/year. The contribution of sludge adsorbed azo dyes to the total amount of azo dyes in landfills may be calculated on the basis of the equation shown below:

Sludge amount to landfills = release ´ adsorption factor ´ fraction to landfills.

In a worst case scenario, the total contribution (processing + use) may be 13.5 tonnes per year, and in a best case scenario 6.70 tonnes per year, which is approximately 6% and 3%, respectively of the total amount of dyes deposited in landfills.

Thus, the total release of azo dyes to landfills may be estimated to approximately 240 tonnes per year in worst case and 233 tonnes per year in best case.

Assuming that the processing industry carries out waste water treatment, the PECagri sludge is reduced to the range of 0.2 to 0.3 mg/kg. The contribution from the use phase is unchanged with 0.25 to 0.5 mg/kg soil. Due to the lack of monitoring data, it is not possible to validate the calculated PECs. However, these concentrations are, compared to a worst case level of 1 mg/kg (w/w of dry soil) reported by Brown and Anliker (1988), lower.

The fate of products containing dyes released to landfills is uncertain, but there may be a potential release of dyes to soil from this compartment.

Ecotoxicity

Aquatic compartment

Azo dyes

Reactive Black 5 (diazo) has a low toxic potential in aquatic organisms (fish LC50 100-500 mg/l; bacteria EC50 > 2,000 mg/l) as well as the hydrolysed dye (fish LC50 > 500 mg/l; Daphnia magna EC50 (48h) > 128 mg/l) (Hunger & Jung, 1991, IUCLID). Very little information is available on the aquatic toxicity of the hydrolysed reactive dyes, but their loss of ability to react with various groups of vital organic materials, such as proteins and DNA, reduces the potential hazard considerably (ETAD, 1991).

Spencer (1984) has examined the effect of Aquashade (a mixture of Acid Blue 9 and Acid Yellow 23*) on the oxygen consumption of the crayfish Orconectes propinquus and has not found any effect at a concentration of 1 mg/l at an exposure of five days.

A survey of available fish toxicity data on over 3,000 commercially available organic dyes by ETAD member companies indicated that about 98% have a LC50 greater than 1 mg/l, a concentration at which colouring of a river normally would be observable. The remaining 2% were acute toxic (LC50 < 1 mg/l). The latter, consisted of 27 different chemical structures including four Acid dyes, sixteen Basic dyes of which 10 were of the triphenylmethane (not azo) type. In only one case, the LC50 was as low as 0.01 mg/l (Clarke & Anliker, 1980). The LC50 for 59% was more than 100 mg/l (Anliker, 1986), indicating that 41% of the organic dyes are potentially toxic or toxic at levels in the range of 1 to 100 mg/l.

Many acid dyes, including azo dyes, exhibit high toxicity to fish but do not significantly inhibit algal growth (Clarke & Anliker, 1980).

Zhang et al. (1995) showed that azo dyes competitively inhibit COD utilisation or respiratory rates of biofilms at concentrations of 10 mg/l of Acid Orange 14. However, the inhibition effect was much less significant in biofilms, compared to a suspended activated sludge system. Furthermore, the results indicated that the aerobically non-biodegradable dyes, Acid Orange 10* and 14, were more toxic compared to biodegradable dyes such as Acid Orange 7* and 8.

Brown et al. (1981) reported the results of a study of possible inhibitory effects of dyes on aerobic waste water bacteria measured as respiratory rate. They tested both acid, direct, disperse, reactive, basic, vat, solvent and mordant dyes. The study indicated that 18 out of 202 dyes showed an IC50 less than 100 mg/l, including three dyes with an IC50 between 1-10 mg/l. These 18 dyes were all basic dyes. Unfortunately it was a mixture of chemical classes of dyestuffs, including azo dyes, so it is not possible to relate the results directly to specific azo compounds.

ICI found no adverse effects on the carp (Cyprinus carpio) exposed to less than 10 mg/l of 30 water soluble (ionic) and 12 disperse dyes for 8 weeks (Brown, 1987).

Dyes in the aquatic environment were reported to affect microbial populations and their activities. Azo dyes such as Basic Brown 4, Direct Brown 95*, Direct Black 80, Mordant Black 11, Acid Black 52, Direct Red 81* and Direct Yellow 106 were inhibitory to microbial oxidation processes in both activated sludge and stream water. The inhibition by the basic dyes were stronger than the inhibition by acid dyes when the pH was above the isoelectric point of the micro-organism. The inhibition was weakened by introduction of the functional groups methyl, nitro, sulpho or acid to the azo dye or by replacement of the benzene ring with a naphthalene ring. However, introduction of chlorine or bromine strengthened the observed inhibition (Chung & Stevens, 1993). The IC50 was not stated.

In an ADMI (American Dye Manufacturers Institute) study, the toxic effects of 56 selected dyes to the green alga Selenastrum capricornutum were examined. The growth of the algae was assessed after 7 and 14 days in the presence of 1 and 10 mg/l of dyes. 15 dyes (27%) strongly inhibited growth at a test concentration of 1 mg/l after 7 days of incubation (Brown & Anliker, 1988).

The following short term test results are available from a study by ETAD and presented on a seminar in 1992 (ETAD, 1992b). ETAD carried out an investigation of 47 dyes of different chemical dye classes. Even though the specific amount of azo dyes in the investigation is not stated, the results are shown in Table 5.18, Table 5.19, Table 5.20, Table 5.21, Table 5.22 and Table 5.23 below, in order to gain insight to the toxicity of the different chemical (technical) groups of dyes.

Table 5.18

Toxicity of Acid dyes, a total of 11 (ETAD, 1992b).

Syre farvestoffers toksicitet, ialt 11 (ETAD, 1992b).

Test organism End point

No. of

Toxicity mg/l

 

results

<1

1-10

10-100 >100
Zebra fish 96 hr LC50

11

0

2

3

6

Daphnia Magna 48 hr EC50

9

0

0

6

3

Alga 72 hr EC50

9

2

3

3

1

Bacteria IC50

11

0

0

0

11

Table 5.19

Toxicity of Basic dyes, a total of 6 (ETAD, 1992b).

Basiske farvestoffers toksicitet, ialt 6 (ETAD,1992b).

Test organism End point

No. of

Toxicity mg/l

    results

<1

 

1-10

10-100

>100

Zebra fish 96 hr LC50

6

0

 

3

3

0

Daphnia Magna 48 hr EC50

6

5

 

0

1

0

Alga 72 hr EC50

6

2

 

4

0

0

Bacteria IC50

6

0

 

3

2

1

Table 5.20

Toxicity of Hydrolysed Reactive dyes, a total of 8 (ETAD, 1992b).

Hydrolyserede reaktive farvestoffers toksicitet, ialt 8 (ETAD, 1992b).

Test organism End Point

No. of

Toxicity mg/l

       
    results

<11

1

1-10

10-100

>100

Zebra fish 96 hr LC50

8

0

 

0

0

8

Daphnia Magna 48 hr EC50

8

0

 

0

0

8

Alga 72 hr EC50

8

0

 

0

71

1

Bacteria IC50

8

0

 

0

0

8

1 Alga results are >10 mg/l.

Table 5.21

Toxicity of Direct dyes, a total of 7 (ETAD, 1992b).

Direkte farvestoffers toksicitet, ialt 7 (ETAD, 1992b).

Test organism End point

No. of

Toxicity mg/l

   
    results

<1

 

1-10

10-100

>100

Zebra fish 96 hr LC50

7

0

 

0

0

7

Daphnia Magna 48 hr EC50

7

0

 

0

0

7

Alga 72 hr EC50

7

0

 

3

2

2

Bacteria IC50

7

0

 

0

0

7

Table 5.22

Toxicity of Disperse dyes, a total of 11 (ETAD, 1992b).

Disperse farvestoffers toksicitet, ialt 11 (ETAD, 1992b).

Test organism End point

No. of

Toxicity mg/l

       
    results

<1

 

1-10

10-100

>100

Zebra fish 96 hr LC50

11

0

 

0

2

9

Daphnia Magna 48 hr EC50

10

0

 

2

2

6

Alga 72 hr EC50

8

3

 

3

2 (3)1

0

Bacteria IC50

11

0

 

0

0

11

1 3 alga results are >10 mg/l.

Table 5.23

Toxicity of Mordant dyes, a total of 3 (ETAD, 1992b).

Mordante farvestoffers toksicitet, ialt 3(ETAD, 1992b).

Test organism End point

No. of

Toxicity mg/l

       
    results

<1

 

1-10

10-100

>100

Zebra fish 96 hr LC50

3

0

 

0

0

3

Daphnia Magna 48 hr EC50

1

0

 

0

0

1

Alga 72 hr EC50

3

3

 

0

0

0

Bacteria IC50

3

0

 

0

0

3

It is indicated that the bacteria are less susceptible to the different classes of dyes compared to other test organisms. Among the tested dyes, the bacteria were only susceptible to basic dyes at concentrations below 100 mg/l, which is in agreement with the findings of Brown et al. (1981) and Chung and Stevens (1993).

From the tables it is indicated that the zebra fish is susceptible to (in declining order) basic dyes > acid dyes > disperse dyes at a level less than 100 mg/l. For the other chemical classes, hydrolysed reactive, direct and mordant dyes, the LC50 is above 100 mg/l. The susceptibility to acid and basic dyes for fish is in agreement with the findings of Clarke and Anliker, 1980.

The susceptibility of Daphnia resembles that of the zebra fish, but the order is different, basic > disperse > acid. The remaining chemical classes all show a LC50 above 100 mg/l. The study confirms the findings reported by Hunger and Jung (1991) and IUCLID that the reactive dyes and hydrolysed reactive dyes have a low toxic potential in aquatic organisms.

The alga is apparently the most susceptible organism, because it for all the tested dyes showed a susceptibility to the dye below 100 mg/l. The susceptibility was in the following declining order Mordant > Basic/acid/-disperse > direct > hydrolysed reactive dyes.

Based on literature and database studies, it was possible to obtain results of the LC50 for some of the azo dyes in use in Denmark, but in general there are only a few data available on effects through the normal sources (AQUIRE, IUCLID, HSDB, MITI, etc.). Furthermore, only data on various fish species were obtained, and it was not possible to obtain data on the basic, mordant and the disperse dyes which are used in Denmark.

In Table 5.24, the lowest effect concentrations for azo dyes used in Denmark are presented. The data indicate that various fish species are susceptible to acid and direct dyes at a level between 1 to 10 mg/l. The susceptibility regarding solvent dyes is in one instance below 1 mg/l. It is not known, if some of these azo dyes were included in the study by ETAD (1992b).

Table 5.24

The lowest effect concentrations for azo dyes used in Denmark.

Laveste effektkoncentrationer for azofarvestoffer anvendt i Danmark.

C.I. name Fish Effect

Conc.

  Organism  

mg/l

Acid Blue 1131 Pimephales promelas LC50, 96 h

4

Acid Red 1142 Cyprinus carpio LC50, 48 h

4

Direct Blue 141 Oncorhynchus mykiss LC50, 24 h

6

  Oncorhynchus tschawytchia LC50, 24 h

6

  Ptychocheilus oregonensis LC50, 24 h

10

Solvent Yellow 11 Oryzias latipes LC50, 58 h

0.7

Solvent Yellow 31 Leuciscus delineatus EC50, 96 h

2

1 AQUIRE.

2 MITI.

In addition to the figures shown in Table 5.24, one fish species (Oryzias latipes), exposed 48 hours to Acid Yellow 36, had a LC50 of 68 mg/l. For the remaining dyes, amongst them 5 acid, 6 direct and 2 solvent dyes, the LC50 was above and well above 100 mg/l. Apparently, the different almost exclusively fish species show very variable susceptibility. For further details, see Appendix 2.

Azo compounds

At exposure levels of 2,500 ppm of the azo compound 3,3´, 4,4´-tetrachloroazobenzene on diet, the mortality of the Japanese Medaka (Oryzias latipes) was significantly higher compared to the control group (Allison & Morita, 1995b). On the other hand detrital exposure levels of 2,500 ppm of the same compound did not appear to cause any harmful effects towards the aquatic snail (Indoplanorbis exustus) (Allison & Morita, 1995a).

Metabolites

Couch and Harshbarger (1985) presented an overview of the effects of carcinogenic agents on aquatic animals in experimental studies. They found reports on the effects of aminoazotoluene on fish, adult guppy and adult Medaka at dietary exposure levels of 120 mg/l and 600 mg/l, respectively. Further neoplasm in the liver was induced within 12 weeks and 24 weeks, respectively. The argument of the authors was that in the environment the susceptibility to xenobiotics may differ among different species. Subsequently, proliferation and cellular disorder are neoplasms, which may be caused by xenobiotics, viruses or an interaction of both.

Hermens et al. (1990) investigated the influence on enzyme induction (MFO P450) on the acute toxicity (96-hr LC50) of 4-chloroaniline (p-chloroaniline) to the rainbow trout (Salmo gairdneri). The 95-hr LC50 was from 11.0 to 14.0 mg/l, and the results showed no significant difference between prior induced trout (50 mg/kg Aroclor 1254) and not induced trout, suggesting that metabolic activation does not necessarily play a role in the acute toxicity of aromatic amines to fish.

Metabolic activation of aromatic amines has been shown in the phyla: Mollusca, Crustacea and Echinodermata, e.g. Mytilus edulis, Mytilus galloprovincialis, Carcinus maenas, Asterias rubens, resulting in mutagenicity to Salmonella typhimurium (Marsh et al., 1992).

Dumpert (1987) showed that p-chloroaniline has a lethal effect on the embryo of Xenopus laevis at a concentration of 100 mg/l. Its development is inhibited (teratogenic) at concentrations of 1 and 10 mg/l, respectively.

In a study of bacterial growth kinetics to in vitro toxicity assessment of substituted phenols and anilines, Nendza and Seydel (1990) demonstrated that these compounds were inhibitory, and that the toxic action was probably caused by damage to the bacterial cells. This was documented by decrease in growth rate and in change of the Na+/K+ ratio with an increase in the Na+ and a decrease in the K+ concentrations. Furthermore, the authors found a good agreement between growth kinetics of E.coli and fish tests (guppy and zebra fish) for phenols - a linear relationship between log 1/LD50 guppy and log 1/I50 E.coli.

p-aminoazobenzene (10.23 mg/l) was by Zissi and Lyberatos (1996) found to result in a decrease of 15% in the specific growth rate of Bacillus subtilis.

The lowest effect concentrations found for the restricted aromatic amines are presented in 5.25 below. No data were obtained on the naphthalene based amines.

Table 5.25

The lowest effect concentrations for some of the metabolites.

Laveste effektkoncentrationer for visse nedbrydningsprodukter.

Name Organism

Effect

Conc., mg/l

4-chloroaniline Daphnia magna EC50, 24 h

0.061

  Lepomis macrochirus LC50, 96 h

2.01

  Pimephales promelas LC50, 96 h

121

  Salmo gairdneri LC50, 96 h

141

  Xenopus laevis Teratogen

11

4- aminobenzene Oryzias latipes, juv LC50, 24 h

1.72

  Oryzias latipes, juv LC50, 48h

0.72

o-anisidine Daphnia magna EC50, 48 h

6.83

  Poecilia reticulata EC50, 336 h

18

benzidine Limoria lignorum, adult LC100, 18 h

>0.052

  Oryzias latipes, juv LC50, 24 h

16.02

  Oryzias latipes, juv LC50, 48 h

10.52

1 Dumpert (1987).

2 ECDIN.

3 Federal Ministry of the Environment, Youth and Family, Austria 1997.

As shown in Table 5.25, it has been reported that benzidine and 4-aminobenzene are acute toxic (LC50 < 1 mg/l) to some crustaceans and juvenile fish. The EC50 for Daphnia magna is as low as 0.06 mg/l for

4-chloroaniline. In general, the LC50 of 4-chloroaniline for various fish species is in the range of 12 to 46 mg/l which indicates potential toxicity. o-anisidine has a LC50 (336 h) of 165 ppm towards adult Poecilia reticulata. o-toluidine has a LC50 (336 h) of 81 ppm towards juvenile Poecilia reticulata. For further details, see Appendix 3.

Summary

No specific data were obtained on basic, reactive, mordant and disperse dyes for any of the dyes encompassed in the present survey.

But it may be concluded that some of the acid and basic dyes are acute toxic to toxic to aquatic organisms (fish, crustaceans, algae and bacteria), which also applies for at least some of the direct dyes, e.g. Direct Blue 14*. Reactive dyes (Reactive Black 5) generally have very high effect concentration levels (>100 mg/l) and are not considered to be toxic to aquatic organisms.

Furthermore, it is indicated that the non-ionic (disperse, mordant and solvent) dyes are toxic and potentially toxic. Solvent dyes may even be acute toxic to aquatic organisms. The mordant dyes may, according to the present findings, not exhibit any toxicity at levels below 100 mg/l.

Algae are generally susceptible to dyes, but the inhibitory effect is thought to be related to light inhibition at high dye concentrations, rather than a direct inhibitory effect of the dyes. According to ETAD (1994), this effect may account for up to 50% of the inhibition observed.

The effects of the substitutional pattern of the dyes are inconclusive, but it has been suggested that introduction of the functional groups; methyl, nitro, sulpho or acid, weakens the inhibition of bacteria, whereas introduction of chlorine and bromine strengthens the inhibition.

In general, it should be noted that toxicity data of chronic low-level exposures for most of the commercial dyes and their derivatives are lacking.

It is indicated, in general, that the effects of the metabolites to aquatic organisms, except for algae, are at levels where potential toxicity is recognised (LC50 < 100 mg/l). This applies for all of the three groups: anilines, benzidines and toluidines. No data were obtained for the naphthalenes.

Anilines and benzidines are both acute toxic and toxic depending on the specific species. The anilines seem to be more toxic to Oryzias latipes juv than benzidine. The findings of the toluidines indicate potential toxicity for various aquatic organisms.

PNEC - Dyes

Applying an assessment factor of 100 on the EC50 from respiration inhibition test (Table 5.20), the following PNEC is derived in accordance with TGD Part II, section 3.4:

PNECstp is in the range of 10 : g/l to 100 : g/l.

It should be noted, however, that it is not known if the observed effect is caused by azo dyes or other dye types. But the significance of possible inhibitory effects of azo dyes to the bacteria in the sewage treatment plant is of great importance, therefore, the estimate of PNECstp is included.

Short term data from each of the three trophic levels (alga, fish, daphnia) of the base set are available. Hence, according to TGD Part II, section 3.3.1 an assessment factor of 1,000 is applied at the lowest L(E)C50. However, as stated above it is not known, if the observed effect is caused by azo dyes in the case of algae and daphnia, cf. Table 5.18 and Table 5.19. But the lowest observed effect is observed for fish (Table 5.24) with a LC50 of 0.7 mg/l for Oryzias latipes, arriving at a PNEC of:

PNECaquatic organisms = 0.7 : g/l.

Metabolites

No data were obtained on bacterial inhibition of the metabolites.

Data of two trophic levels were obtained for the metabolites. The lowest observed effect is found in daphnia (Table 5.25) with a LC50 of 0.06 mg/l for Daphnia magna, arriving at a PNEC of :

PNECaquatic organisms = 0.06 : g/l.

Atmosphere

No data were obtained on atmospheric exposure.

Terrestrial compartment

ETAD has organised a study of the possible effects of dyes on plant germination and growth. Four dyes were used and among them an acid dye of the azo type (C.I. 13155). All four dyes were incorporated into a seed compost at concentrations of 1, 10, 100 and 1,000 mg/l and germination and growth of three plant species (sorghum, sunflower, and soya) were assessed. No effects were observed on seed germination. With respect to the growth rate, there was no observed effects at a concentration of 100 mg/l. At a level of 1,000 mg/l, however, there was a variable growth depending on the dye and the species of the particular plants. After a growth period of 21days, the plant foliage was analysed. At the 1,000 mg/kg soil level the dyestuffs, among them the azo (C.I. 13155), were just detectable in the plant foliage (max. 2 mg/kg) (Brown & Anliker, 1988).

Chung et al. (1997) found out that growth of the soil living nitrogen-fixing bacterium Azotobacter vinelandii is inhibited by p-phenylene- diamine and 2,5-diaminotoluene, which are derivatives after azo reduc-tion of e.g. Basic Brown (C.I. 21010) and Direct Black 80. The nitroge- nous activity was also significantly inhibited at a concentration of 50 : g/ml. p-phenylene- diamine was found to be inhibitory to the growth of other common aquatic and soil bacteria.

PNEC - terrestrial

According to the TGD Part II section 3.6.2.2, an assessment factor of 1,000 should be applied for L(E)C50 short-termed toxicity tests for soil. Brown and Anliker (1988) have reported effects at a level of 1,000 mg/kg for plants, indicating a PNEC of :

PNECsoil = 1 mg/kg.

Risk characterisation

The PEC/PNEC ratios which can be derived with the available data are shown in Table 5.26.

Table 5.26

PEC/PNEC ratios for the aquatic and terrestrial compartments.

PEC/PNEC forhold for vand- og jordmiljø.

Compartment Site  

PEC (mg/l)

PEC/PNEC

     

Worst

Best

Worst

Best

STP Sludge Processing        
    Textile

8.24

4.11

82.4

41.1

    Leather

4.70

2.45

47.0

24.5

    Use        
    Textile

0.68

0.34

6.8

3.4

Aquatic STP effluent Processing        
    Textile

1.44

0.48

2057

685

    Leather

0.82

0.27

1171

386

    Use        
    Textile

0.12

0.04

171

57

  Sediment Processing        
    Textile

0.11

0.02

1.1

0.2

    Leather

0.06

0.01

0.6

0.1

    Use        
    Textile

0.008

0.002

0.1

0.02

  Surf. water Processing        
    Textile

0.14

0.048

200

69

    Leather

0.08

0.027

114

39

    Use        
    Textile

0.012

0.004

17

8

Terrestrial Agr. sludge Processing        
    Textile

0.484

0.242

0.5

0.2

    Leather

0.007

0.003

<0.01

<0.01

    Use        
    Textile

0.498

0.249

0.5

0.3

For substances with a PEC/PNEC ratio of less than 1 there is, according to TGD, no need for further testing and risk reduction measures beyond those which are already being applied. A ratio higher than 1, however, indicates a need for further information and/or testing or even a need for limiting the environmental risks.

It is indicated that there is a need for further testing and information with regards to the concentration of dye in the aquatic compartment, except for the sediment, because the PEC/PNEC is higher than 1. Whereas the PEC/PNEC ratios for the terrestrial compartment indicate, that there is no need for further testing and/or information.

Furthermore, it is indicated that there is a need of further information with regards to the concentration of dyes in the STP, because the PEC/PNEC is well above 1.

With reference to the assumptions and recalculation of the PECs, it is indicated that the PEC/PNEC ratios presented in Table 5.26 are to high.

Recalculation of the PEC/PNEC ratios indicates:

PEC/PNECsludge, stp
3.4 to 5 (>1)
PEC/PNECeffluent, stp
34 to 1,234 (>>1)
PEC/PNECsurface water
3 to 123 (>1)
PEC/PNECsediment
0.01 to 0.9 (<1)
PEC/PNECagri sludge
0.5 to 0.8 (<1)

Summary

The survey indicates that there is a need for further information and testing in order to assess the environmental risk in the STP, the STP effluent and surface water, whereas the releases associated with sludge application in agricultural soil not seem to present any immediate concern.

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