Environmental and Health Assessment of Substances in Household Detergents and Cosmetic Detergent Products
Eutrophication and associated problems have received considerable attention during many years, particularly with respect to the effects on freshwater lakes tending to be phosphorus limited (Lee et al. 1978). In order to reduce the phosphorus content of municipal sewage, voluntary and statutory restrictions have been introduced to limit the use of the detergent builder sodium tripolyphosphate (TPP). The release of complexing agents into the environment may affect the distribution and partitioning of metals in soils, sediments and sludge. Complexing agents may potentially cause active desorption of trace metals from particulate matter or interfere with natural sorption processes. Remobilization of metals has expecially been examined for EDTA and NTA (see Sections 7.6 and 7.7) although this is probably a general effect of complexing agents.
From 1947 until the late eighties sodium tripolyphosphate (TPP) was used almost exclusively as the complexing agent in detergents due to its multifunctional contribution to washing and cleaning processes. Other complexing agents like phosphonates, polycarboxylates and zeolite have now partially replaced phosphates in household detergents. However, in Denmark the strategy is to establish phosphorus removal processes at the major wastewater treatment plants a goal which generally speaking has been fulfilled and TTP is still used in many household detergents. Phosphates serve many functions in detergent products. It removes hardness, reduces surfactant use, improves emulsification and dispersion, prevents re-deposition, and controls alkalinity.
Detergents containing phosphorus contribute together with other sources of phosphorus to the eutrophication of many fresh waters. Algae are the first step in the food chain and a number of factors are needed to promote their growth. These factors are sunlight for photosynthesis, temperature, certain water conditions (turbulence) and nutrients like carbon, nitrogen and phosphorus. Typical plant organic matter of aquatic algae and macrophytes contain phosphorus, nitrogen and carbon in approximately the ratios:
Thus broadly speaking, if one of the above mentioned elements is limiting and all other elements are present in excess of physical needs, phosphorus can theoretically generate its weight 500 times in algae, nitrogen 71 times and carbon 12 times in algae (Wetzel 1983).
Whereas the primary production in marine waters is mainly nitrogen limited, freshwaters are considered to be phosphorus limited. A large part of the sewage effluents in many countries is released untreated into freshwater recipients, and here the use of phosphorus as complexing agents is still an environmental concern.
Toxicokinetics and acute toxicity
Polyphosphates are hydrolyzed into smaller units (orthophosphates) in the gut before absorption, which may induce a metabolic acidosis (Gosselin et al. 1984). The orthophosphates are excreted in the urine (HSDB 1998). Ingested diphosphate is readily converted to monophosphate. No diphosphate was found in faeces or urine of rats treated with diets containing up to 5% tetrasodium diphosphate. Diphosphate was almost completely absorbed by the gut and excreted as monophosphate in the urine (IPCS 1982). The acute toxicity of polyphosphonates is low as the lowest LD50 after oral administration is > 1,000 mg/kg body weight (IPCS 1982; ACGIH 1991).
Skin and eye irritation
The most important human health effect, which may be caused by the use of phosphates in household detergents, is the potential irritation to skin and eyes. Polyphosphates are moderately irritating to skin and mucous membrane (Merck 1989). Polyphosphates can be irritating because of their alkalinity. A 1% aqueous solution of TTP has a pH of 9.8 and the pH of concentrated solutions is about 10.5 (Gosselin et al. 1984). Acute studies with tetrasodium diphosphate show that direct contact causes severe irritation and corneal injury in the rabbit eyes and that it may be irritating to skin (ACGIH 1991).
Mutagenicity and carcinogenecity
No mutagenic potential was observed when TTP was tested in a Salmonella/microsome assay (Ames test) and in a chromosomal aberration assay in vitro using a Chinese hamster fibroblast cell line (Ishidate et al. 1984). Tetrasodium pyrophosphate was not mutagenic in an in vitro assay using S. cerevisiae strains and S. typhimurium strains with and without the addition of mammalian metabolic activation preparations (IPCS 1982).
Sodium tripolyphosphate showed no maternal toxicity or teratogenic effects at dose levels up to 238 mg/kg body weight in mice and 40 mg/kg in rats (IPCS 1982). Reproduction studies in three generations of rats on diets with 0.5% TTP were performed. TTP had no effects on fertility or litter size, or on growth or survival on offspring (Hodge 1964). Tetrasodium diphosphate showed no maternal toxicity or teratogenic effects at dose levels up to 130 mg/kg body weight in mice and 238 mg/kg in rats (IPCS 1982).
Polyphosphates are of low toxicity. No tests on sensitization and carcinogenicity were available. Polyphosphates are not included in Annex 1 of list of dangerous substances of Council Directive 67/548/EEC.
Phosphonate compounds containing more than one phosphonate group are effective sequestrants and possess other useful properties such as high water solubility, chemical stability, bleach stabilizing effects, and the ability to prevent precipitation of calcium salts at substoichiometric concentrations.
Phosphonates are characterized by the presence of one or more C-PO3-H2 groups. Most phosphonates are synthesized from phosphorous acid by reaction with formaldehyde and either ammonia or amines.
An example of a phosphonate synthesized by reaction with ammonia is:
Examples of phosphonates synthesized by reaction with amines are:
1-Hydroxy ethane diphosphonic acid (HEDP; CAS No. 2809-21-4) is formed from PCl3 and acetic acid (Gledhill and Feijtel 1992).
A large percentage of European phosphonate consumption occurs in detergents and, thus, phosphonates are continuously released to the environment in Europe. At present sufficiently sensitive analytical methods for measuring phosphonates are unavailable and environmental concentrations are predicted from models. According to the model simulations the maximum phosphonate levels in aquatic environments are expected to be < 30 m g/l. If partitioning to sediments (100:1) and limited photo- and biodegradation are assumed, the average phosphonate concentrations in European streams are predicted to be in the order of 0.25 m g/l (Gledhill and Feijtel 1992).
A variety of natural and synthetic chemicals contain a C-P bond. The C-P bond provides the molecule stability and a relatively high resistance to chemical, photolytic and thermal decomposition. Phosphonates seem to be recognized by bacteria only as a possible P-source, which may explain poor results in standard biodegradation tests. However, several laboratory studies report phosphonate degradation by pure microbial cultures when supplied as the sole source of phosphorus (Gledhill and Feijtel 1992). Orthophosphate has been found to suppress phosphonate utilisation in many microrganisms. Thus organisms preferentially use inorganic phosphate, which may explain the low biodegradability of phosphonates in synthetic test media and natural sewage systems.
Numerous studies have shown that little, if any, primary or ultimate biodegradation occurs for any phosphonate product in standard biodegradation tests such as the OECD screening test, BOD20 test, sapromat test and closed bottle test. Phosphonates may serve as a carbon source when present at very high concentrations and, e.g., DOC removals of 23-33% of HEDP and ATMP have been observed in a Zahn-Wellens test (Gledhill and Feijtel 1992).
Reports of anaerobic biodegradability are sparse. For HEDP and ATMP less than 4% of the 14C-labelled phosphonate carbon was converted to 14CO2 and 14CH4 in a model digestor (Gledhill and Feijtel 1992).
As expected for highly water-soluble substances, the log Kow values for phosphonates are low (ATMP: -3.53; HEDP: 3.49; EDTMP: 4.10; HDTMP: 4.43; DTMP: -3.40). The potential for bioaccumulation of phosphonates in aquatic organisms is therefore expected to be low as well. Experimental bioconcentration studies with zebra fish have been conducted with radiolabelled ATMP and HEDP. For both substances, the BCF values determined after 4-6 weeks of exposure were less than 24.
Investigation of the effects of phosphonates in alga bioassays is quite complex as the alga medium contains a precise level of micro nutrients which are held in solution by a chelator, EDTA. The introduction of an additional chelator, such as a phosphonate, may indirectly either inhibit or stimulate alga growth. The phosphonate may bind essential metals (indirect inhibition) or it may release additional phosphorus via photodegradation (indirect stimulation). Cell counts were performed at day 4 and day 14 during a toxicity study with Selenastrum capricornutum. The day 4 results indicate EC50 values for the examined phosphonates between 0.4 and 30 mg/l, with EDTMP being the most toxic (Table 7.1). For HEDP, EDTMP and DTPMP the EC50 values measured on day 14 were lower than the values measured on day 4. Initial concentrations of phosphonates may have chelated some essential micronutrients for alga growth, thus resulting in the low EC50 values observed on day 4. In the time period from day 4 to day 14, HEDP, EDTMP and DTPMP may have photodegraded to release these nutrients plus additional phosphorus, which resulted in observed growth stimulation and thus the lower EC50 value observed on day 14. For the reasons described above, the apparent toxicity of phosphonates to algae cannot be regarded as a reliable indication of the toxicity of phosphonates in the aquatic environment.
Generally, the acute EC/LC50 values for phosphonates towards invertebrates are well above 100 mg/l. One exception is the Eastern oyster for which acute LC50 values below 100 mg/l are found (Table 7.2).
As also noted for invertebrates, the LC50 values for phosphonates are well above 100 mg/l. The aquatic toxicity data obtained in long-term studies with fish are not markedly different from the data from short-term studies (96 hours). This indicates that phosphonates do not accumulate and that the maximum toxicity is obtained in short term tests (Table 7.3).
Toxicokinetics and acute toxicity
The intestinal absorption and kinetics of 32Plabelled HEDP have been studied in man. After oral administration 70-90% of the administered dose was found in faeces after 6 days. HEDP was poorly absorbed (Caniggia and Gennari 1977). When 32P-labelled HEDP was given intravenously 35-50% of the administered dose was excreted in the urine after 6 days. No metabolism occurred (Caniggia and Gennari 1977). 14C-labelled EDTMP was poorly absorbed from the gastrointestinal tract and most of the absorbed dose was rapidly excreted by the kidneys or sequestered in bone. EDTMP is not metabolized as the entire radioactivity in the urine was identified as EDTMP (Calvin et al. 1988). No data on percutaneous absorption was available.
Phosphonates show low oral and dermal toxicity (Table 7.4).
Concentrated solutions of ATMP and HEDP have pH values of about 2.1. Only moderately skin and eye irritation have been seen (SFT 1991). In Guinea pig maximization test 1,2,4-butantricarboxylic acid, 2-phosphono, tetrasodium salt in a 32% solution was not sensitizing (IUCLID 2000). ATMP, HEDP and EDTMP did not show sensitizing effects (SFT 1991).
Mutagenicity and carcinogenecity
EDTMP was tested for genotoxicity in the Ames, mouse lymphoma, unscheduled DNA synthesis and in vivo cytogenetics assays. No mutagenic activity was seen in any of the assays (Calvin et al. 1988). HEDP showed no mutagenicity in microsome test with Salmonella typhimurium and mouse lymphoma assays (SFT 1991). A 50% solution of 1,2,4-butantricarboxylic acid, 2-phosphono was tested according to Guideline 474 "Genetic toxicology: Micronucleus Test " as a single oral administration in mice. No mutagenic effects were seen (IUCLID 2000). Rats were fed EDTMP in the diet over a 2 year period. The dose was up to 100 mg/kg/day. No carcinogenic potential was seen (Calvin et al. 1988).
Rabbits were given HEDP by gavage in the doses 25, 50 and 100 mg/kg/day from day 2 to 16 of gestation. No differences between the controls and the treated animals were seen with respect to teratogenicity and maternal toxicity (Nolen and Buehler 1971). A 49% solution of 1,2,4-butantricarboxylic acid, 2-phosphono was given orally to rats. The doses were 0, 100, 300 or 1,000 mg/kg and they were given from day 6 to 15 of gestation according to Guideline 414 "teratogenicity". No teratogenicity, embryotoxicity or maternal toxicity were seen (IUCLID 2000).
Phosphonates show no sensitizing, mutagenic or reproductive effects. Low acute oral and dermal toxicity is seen. Phosphonates are not included in Annex 1 of list of dangerous substances of Council Directive 67/548/EEC.
Polycarboxylates used in washing powders and detergents are homopolymers of acrylic acid or copolymers of acrylic acid and maleic anhydride, generally as sodium salts. Relevant CAS Nos. are: Sodium polyacrylate (9003-04-7), polyacrylic acid (9003-01-4), and acrylic acid polymers with maleic anhydride, sodium salt (52255-49-9). The various polycarboxylates are distinguished by the monomers used for their preparation, acrylic acid (AA) and maleic anhydride (MA), and their mass-average molar mass or molecular weight (MW). The polymers are designated by codes of the corresponding abbreviations, P(AA) for polyacrylic acid, and P(AA-MA) for the copolymer of acrylic acid and maleic anhydride, to which the numerical value of MW is suffixed.
As a consequence of the reduction of phosphate content in detergents, the concentrations of free calcium and magnesium rise in the washing water. The metal ions tend to form precipitates with hard water and some detergent components. Polycarboxylates inhibit the crystal growth of inorganic precipitates so that these salts remain in suspension and do not precipitate onto textile fabrics
Due to their major use in detergents, the main route for the emission of polycarboxylates to the environment is via domestic sewage treatment plants to surface waters receiving the treated effluents.
Polycarboxylates are removed from sewage water by physico-chemical processes such as sorption onto particulate matter and precipitation, which implies that the polycarboxylates will partition into the sludge. Sewage sludge is frequently stabilised by anaerobic digestion and subsequently used as fertilizer in agriculture. Therefore, degradation and elimination processes in sewage treatment plants, surface waters and soils are of main interest.
Polycarboxylates are generally not rapidly biodegradable, and no evidence for short-term biodegradation has been obtained when P(AA)3,000-4,000 was evaluated for BOD5, BOD10 and DOC removal in test systems inoculated with effluent from a municipal sewage treatment plant (ECETOC 1993). A respirometric screening test with P(AA-MA)70,000 showed a biodegradability corresponding to < 14% biodegradation. A number of 14C-labelled P(AA)1,000; 2,000; 4,500; 10,000 and P(AA-MA)12,000; 70,000 were tested in flasks fitted with CO2 absorbers. The polycarboxylates (test concentration: 0.1 and 1 mg/l) were incubated for up to 19 weeks in river water, pre-adapted river water or a mixture of river water and sediment. Mineralisation in river water was < 20% for all polymers tested. The P(AA) were mineralised to a higher degree in pre-adapted river water and river water plus sediments than in river water alone: 63% and 58% for P(AA)1,000 and 15% and 12% for P(AA)10,000). The results for P(AA-MA) were not significantly different in the three test waters and indicate that their degradation is slow (< 20%) under discontinuous test conditions (ECETOC 1993).
A partial biodegradation of polycarboxylates with a molecular weight of 1,000-70,000 has been indicated in tests with activated sludge inoculum. P(AA)1,000 was mineralised to an average extent of 43%, whereas P(AA)2,000 and P(AA-MA)70,000 were mineralized 19% and 15%, respectively (ECETOC 1993).
The fate of radiolabelled P(AA-MA)70,000 has been examined in a sewage-treatment plant model system using pre-adapted sewage sludge. Both by continuous and pulse loading more than 90% of the 14C was recovered in the sludge, while 2-3% remained in the supernatant (ECETOC 1993).
Several studies have shown that the biodegradation of polycarboxylates in soils is poor. E.g., the 14CO2 production was followed for 1 year in a standard soil which was treated with 14C-labelled P(AA-MA)70,000. The total formation of 14CO2 was 4-7% of the added 14C, and it occurred mainly within the first month (ECETOC 1993).
No evidence exists for the biodegradation of high molecular weight polycarboxylates under anoxic conditions. Anaerobic incubation of 14C-labelled P(AA-MA)70,000 in a model digester containing domestic sewage sludge showed that the substance was not mineralized under the applied conditions as 94-95% of the added 14C was sorbed to the sludge particles (ECETOC 1993).
No experimental data are available describing the bioaccumulation potential of polycar-boxylates. However, the molecular weight of polycarboxylates used in laundry detergents is normally between approx. 1,000 and 100,000 and, hence, the bioaccumulation potential of typical commodity chemicals is assumed to be low. Uptake through biological membranes is only anticipated for substances with a molecular weight < 1,000 (OECD 2000).
Inhibitory effects on the growth rate of algae have been observed with Scenedesmus subspicatus, where the 96 h-EC10 was 180 mg/l for P(AA)4,500. The 96 h-EC10 values were 32 mg/l and ³ 200 mg/l for P(AA-MA)70,000 in tests with Scenedesmus subspicatus. A similar low toxicity was seen for P(AA)78,000 as the EC10 (4-14 days) were 82 mg/l for Scenedesmus subspicatus and 30 to more than 1,000 mg/l for Chlorella kessleri (ECETOC 1993).
Crustaceans and fish
Data describing the acute toxicity for daphnids and fish are available for a number of polycarboxylates with different molecular weights. A consistently low toxicity has been observed with LC50 above the highest concentration tested (LC50 > 100 1,000 mg/l) (ECETOC 1993).
Sediment and fish
The acute toxicity of P(AA)4,500 to chironomid larvae was tested in a sediment batch system. After 96 hours, no effects were observed at the highest concentration tested (4,500 mg/kg dry matter) (ECETOC 1993). The acute toxicity of polycarboxylates to earth worms (Eisenia foetida) is also low. For P(AA)4,500 the LC50 was > 1,000 mg/kg soil. The LC0 values reported for P(AA)78,000 and for P(AA-MA)70,000 were 1,000 mg/kg soil and 1,600 mg/kg soil, respectively.
Toxicokinetics and acute toxicity
14C-labelled P(AA-P)2,500 (50% aqueous solution of phosphonated P(AA)) was given to rats by gavage in the concentrations of 25 mg/kg body weight. After 4 days 0.35% of the administered dose was recovered in expired air, 0.47% in the urine and 82-94% was recovered in the faeces. This result indicates a very little absorption from the intestinal tract (ECETOC 1993). In a study of skin penetration of P(AA-P)2,500, only 0.3% was recovered after 2 days in expired air, urine and faeces combined. In general components with a molecular weight > 1,000 have difficulties in penetrating the skin (ECETOC 1993).
The LD50 values by oral administration for rats and mice are over 5 g/kg body weight and by dermal administration for rabbits over 5 g/kg body weight (ECETOC 1993). These values indicate a low acute toxicity by oral and dermal administration.
Skin and eye irritation
Irritation of polycarboxylates has not been observed in man. A 40% active solution of P(AA)7,000 and a 45% solution of P(AA)8,000 were not irritant to the skin of rabbits (ECETOC 1993). When P(AA)1,000 or P(AA)1,200 were applied to the eyes no damage to the cornea or iris was observed. A slight conjunctivae irritation was observed but this cleared within 24 hours after administration. The concentrations were no further specified (ECETOC 1993). P(AA) with different molecular weights were not found to be sensitising (ECETOC 1993).
No serious adverse effects were observed by oral, dermal or pulmonal administration (ECETOC 1993).
Mutagenecity and carcinogenicity
No evidence of mutagenic potential for polycarboxylates P(AA) and P(AA-MA) tested in a variety of genetic tests, such as Ames test, gene mutation in mammalian cells (mouse lymphoma), UDS (unscheduled DNA synthesis) assay and micronucleus test (Thompson et al. 1989). The International Agency for Research on Cancer (IARC) has evaluated polyacrylic acid and the data available to the working group did not permit an evaluation of the carcinogenicity to humans of polyacrylic acid (IARC 1979).
P(AA)90,000 and 4,500 and P(AA-MA)12,000 have been tested. The compounds were administrated by gavage to rats during major organogenesis (on day 6 to 15 of gestation) at dose levels of 500 7,000 mg/kg bw/day. No treatment related adverse effects on foetal development (skeletal abnormalities and soft tissue) were seen (Nolen et al. 1989).
In general P(AA) and P(AA-MA) have a low acute toxicity after oral and dermal administration. No data on carcinogenicity were available. No evidence of a mutagenic and a teratogenic potential has been reported. Some P(AA) were slightly irritating to rabbit eyes. No sensitizing potential has been identified. Polycarboxylates are not included in Annex 1 of list of dangerous substances of Council Directive 67/548/EEC.
Sodium citrates (disodium citrate (CAS No. 144-33-2) and trisodium citrate (CAS No. 68-04-2)) are salts of citric acid. Sodium citrates are widely used in phosphate-free detergents and cleaners. Sodium citrate solutions will exhibit a pH of about 8.5 and are subject to microbial growth. Citrate is a chelating agent for di- and trivalent metal ions.
Sodium citrates are rapidly and ultimately biodegradable under aerobic and anoxic conditions. E.g., sodium citrate attained 90% ThOD in a closed bottle test for ready biodegradability during 30 days (IUCLID 2000).
Due to a low log Kow value of 1.72, sodium citrate is not expected to accumulate in aquatic organisms.
Sodium citrate has a low toxicity towards aquatic organisms (Table 7.5).
Toxicokinetics and acute toxicity
Sodium citrate is oxidized to bicarbonate and excreted in the urine (HSDB 1998). Sodium citrate is a normal human metabolite of carbohydrates in the Krebs cycle (citric acid cycle). It is the glycolytic pathway in which glucose is converted into pyruvate. An LD50-value was found to be 7.1 g/kg body weight after oral administration to mice (Hoyt and Gewanter 1992). This value indicates a low acute toxicity by oral administration.
Skin and eye irritation
Sodium citrates are not irritating to rabbit skin in a test performed according to OECD Guideline 404 (IUCLID 2000).
Sodium citrate is well tolerated by the eye and has proven effective in experimental treatment of cornea injuries caused by alkalines. Sodium citrate reduced the incidence of ulceration and perforation (Grant and Schuman 1993).
Mutagenicity and carcinogenicity
Citrates have shown no mutagenic effects, and no potential carcinogenicity is suspected for citric acid and its salts (Hoyt and Gewanter 1992). Sodium citrate was tested in Salmonella/microsome assay (Ames test) and chromosomal aberrations assay in vitro using a Chinese hamster fibroblast cell line. No mutagenic potential was observed in either test (Ishidate et al. 1984).
Reproductive toxicity Classification
Sodium citrate was negative in teratogenicity studies (Schardein 1993). Based on available information, sodium citrates are considered safe when used in detergents and cleaners. Sodium citrates are not included in Annex 1 of list of dangerous substances of Council Directive 67/548/EEC.
Zeolite (CAS No. 1318-02-1) is an inert, insoluble aluminium silicate, which softens water by ion exchange (Henning et al. 1977).
Removal during wastewater treatment
The removal of zeolite during wastewater treatment is mainly due to sorption to sludge. The removal of zeolite A in primary treatment has been investigated by Carrondo et al. 1981 in static column tests using raw waste water. Removals of 55% and 69% were observed for retention times of one and two hours, respectively. Carrondo et al. 1980 investigated zeolite A removal in the activated sludge process at pilot scale. Assuming a 50% removal during primary treatment, the plant was dosed with zeolite at 15 and 30 mg/l. Average removal rates were 88%, and the removal rate was always greater than 80%.
The toxicity of zeolite towards aquatic organisms is low (Table 7.6).
Toxicokinetics and acute toxicity
The gastrointestinal absorption of silicic acids depends on the degree of polymerisation. The lower the degree of polymerisation, the higher the absorption. Silicic acid monomers were absorbed at a very high rate (Yokoi and Enomoto 1979; IUCLID 2000). In a human stomach model it was found that zeolite Na-A is hydrolyzed to silicates and aluminates. Studies with rats indicate that the silicate is excreted by the urinary system and the aluminate in the faeces (Christophiemk et al. 1992). The LD50-value of zeolites by oral administration for rats is > 5 g/kg body weight (Gloxhuber et al. 1983; IARC 1997).
Skin and eye irritation
Zeolite A is not irritating to rabbit skin according to OECD guideline No. 404 "Acute Dermal Irritation/Corrosion" (IUCLID 2000). In a patch test, a 1% suspension of Zeolite A was exposed to human skin for 24 hours and no irritation was observed (Gloxhuber et al. 1983). Zeolite A is slightly to non-irritating to the eyes (IUCLID 2000). No sensitization potential of zeolites was observed (Gloxhuber et al. 1983; Christophiemk et al. 1992).
No indications of any chronic toxicity have been found.
Mutagenicity and carcinogenicity
Synthetic zeolite A was tested for carcinogenicity in rats by oral administration of 0.6, 6.0 or 60 mg/kg/day via the diet for two years. No increase in tumour incidence was found. No human data are available (Gloxhuber et al. 1993). No indications of any chronic toxic or tumorigenic effects in rats given 1,000 ppm zeolite A (about 50 mg/kg/day) orally in 2 years (Christophiemk et al. 1992). In Salmonella typhimurium strains (Ames test) no mutagenic potential of zeolite A was seen (IUCLID 2000).
Zeolite A was tested for its teratogenic potential in rats and rabbits. The zeolite was given in destilled water by gavage on day 6-15 of gestation for rats and day 6-18 to rabbits. No adverse effects were observed on the dams, the embryo or the fetuses at the doses tested. The highest dose tested was 1,600 mg/kg body weight (Nolen and Dierckman 1983).
Zeolite A is considered non-sensitizing and non-irritating to the skin, but may be slightly eye irritating. No carcinogenic and teratogenic potential has been observed. Very low acute toxicity of zeolites by oral administration is observed. Zeolites are not included in Annex 1 of list of dangerous substances of Council Directive 67/548/EEC.
Ethylenediamine tetraacetate, EDTA, (CAS No. 60-00-4) and EDTA tetrasodium salt (CAS No. 64-02-8) are used at low levels in fabric washing powders as a bleach stabiliser. They are also used in soaps as stabilisers and in some liquid products to enhance the action of preservatives. Besides, EDTA is used in detergents for the industrial and institutional market like, e.g., machine dishwashing agents.
EDTA is a hexadentate chelator capable of combining stoichiometrically with metals. The EDTA-metal interactions depend on metal concentrations, pH, nature of the sediment, concentrations of organics etc. Therefore, it is not possible to give a single value for an EDTA concentration at which no effects on metal remobilization occur. The most preferred EDTA-metal complex is Ni followed by Cu, Zn or Pb, but the formation of these complexes is very dependent on the water-specific conditions. At low EDTA concentrations, nearly all of the EDTA is bound to Ni. With increasing EDTA concentrations other metal ions are complexed successively (EU, Risk Assessment 2000). The metal which forms the most stable EDTA complex is Fe(III), and iron is the most frequent heavy metal in river water. However, studies on the EDTA speciation in surface waters have shown that no major amounts of FeEDTA are present as insoluble Fe(OH)3 and Fe(O)OH are formed. When EDTA is discharged to aquatic environments it has been shown that it will always occur as a metal complex. In German rivers, heavy metal concentrations of approximately 0.5 m mol/l (sum Cd, Cu, Hg, Ni, Pb, Zn) are detected. The stoichiometric EDTA equivalent is about 150 m g/l. In most rivers, the EDTA concentration is lower. Therefore, all EDTA is bound onto actually emitted heavy metal, and there is no free EDTA available to remobilize metals from sediments. Remobilization from the deeper layers is limited by formation of nearly insoluble metal sulfides. Only if the sediments are whirled up during high water flows, a significant increase of heavy metal abundance in the water phase may occur (EU, Risk Assessment 2000).
EDTA is not readily biodegradable in standardized OECD tests, but several lines of evidence suggest that the compound is inherently ultimately biodegradable under aerobic conditions. By use of a pre-adapted inoculum, 10% carbon dioxide evolution and 22% DOC removal were observed in the Sturm test, whereas a higher DOC removal (37%) was attained in the Zahn-Wellens test (Wolf and Gilbert 1992). Other data indicate an interesting relation between pH and the biodegradation of CaNa2EDTA. By using samples from a river, a ditch and a lake as inocula in the closed bottle test, a biodegradation between 60 and 83% was obtained after 49 days at pH 6.5, whereas between 53 and 72% were obtained after 28 days at pH 8.0 (EU Risk Assessment 2000).
No biodegradation of EDTA has been observed under anoxic conditions.
A highly polar, water-soluble compound such as EDTA is not expected to bioaccumulate by partitioning into the lipid component of aquatic organisms. A whole body bioconcentration factor of 1, with a half-life for depuration of 128-242 hours, was observed for bluegill sunfish exposed for 28 days to radiolabelled EDTA (Bishop and Maki 1980). Much information about the influence of EDTA on the accumulation on heavy metals is available. E.g., studies of the influence on Cd accumulation on rainbow trout (Oncorhynchus mykiss) have indicated that EDTA decreases the accumulation of Cd (Pärt and Wikmark 1984).
The toxicity of EDTA to aquatic organisms is dependent on the hardness of the test medium and the pH. This has been shown in experiments with bluegill sunfish (Lepomis macrochirus): The LC50 of EDTA was 61.2 mg/l in very soft water, 401.7 mg/l in medium hard water and 807.3 mg/l in very hard water. At a pH of 3.7 the LC50 was 159 mg/l, 486 mg/l at a pH of 8.9 and 2,340 mg/l at a pH of 7.4 (Wolf and Gilbert 1992).
Information on the chronic toxicity of EDTA towards aquatic organisms is lacking, although it is reported that the NOEC is usually higher than one tenth and almost always higher than one hundredth of the corresponding LC50 (ECETOC 1984).
Toxicokinetics and acute toxicity
Calcium disodium EDTA is poorly absorbed from the gastrointestinal tract in humans with only 2.5% of an oral dose of 3.0 gram being excreted in the urine (Richardson 1992-1994). Studies in rats also indicated that calcium disodium EDTA was poorly absorbed from the gastrointestinal tract. About 8095% of the dose appeared in the faeces after 24 hours. The amount absorbed in 24 hours, determined from the quantity found in the tissues and urine ranged from 218% with most of the values between 2 and 4% (WHO 1998). Low acute toxicity by oral administration is observed. Conversion from the tetrasodium salt to the calcium disodium salt greatly reduced toxicity. The acute toxicity of EDTA is given in Table 7.7.
In a test performed according to OECD Guideline 404 "Acute Dermal Irritation/Corrosion" Tetrasodium EDTA is found non-irritating to the rabbit skin (IUCLID 2000). A 1% aqueous solution of tetrasodium EDTA has a pH of 11.8. Unless first neutralized, EDTA should not be applied to the eyes, because the solutions are alkaline enough to be injurious to the eye (Grant and Schuman 1993). Disodium EDTA has been used therapeutically on the cornea for decontaminating the eye after alkaline splashes and removal of superficial calcific opacities that occur in band keratopathy. EDTA is a common component in many eye drops and contact lens wetting and cleansing solutions (WHO 1998). In normal human eyes, a near neutral 0.1% solution of disodium EDTA applied as eyedrops or as an eye bath causes only mild stinging sensation (Grant and Schuman 1993).
EDTA is not found sensitizing in guinea pig maximization tests (Fisher 1986; IUCLID 2000).
Mutagenicity and carcinogenicity
Trisodium EDTA was tested for its mutagenic potential in Salmonella typhimurium strains and Escherichia coli in laboratories, and no mutagenicity was observed either with or without the S9 metabolic activation system (Dunkel et al. 1985). Trisodium EDTA was tested for its mutagenic potential in the mouse lymphoma cell forward mutation assay, with and without S9 metabolic activation system. No mutagenicity was observed (McGregor et al. 1988).
EDTA induced an increased mutant frequency in a mouse lymphoma assay without metabolic activation at concentrations of 25 and 30 mmol/l. Whether the mutagenic activity of EDTA was due to pH effects which has been suggested earlier is unclear (Wangenheim and Bolcfoldi 1988).
EDTA disodium salt administered 186 mg/kg body weight in mice showed no effects of inducing chromosomal aberrations in mouse spermatogonia, but induced micronuclei in the mouse after treatment of germ cells at the late stages of meiosis (Russo and Lewis 1992).
Zordan et al. (1990) investigated the genetic effects of EDTA disodium salt in the germ cells and the somatic cells in Drosophila melanogaster and mouse. The dosages were 93 and 186 mg/kg body weight. No increase in aneuploidy incidence was seen in bone marrow cells of the mouse and EDTA did not induce increased aneuploidy in spermatocytes of mouse either. EDTA induced aneuploidy in the germ cells of Drosophila but was negative in the somatic cells of Drosophila.
In mouse lymphoma cells DNA-strand breaks were measured in vitro without metabolic activation. There was a clear evidence of DNA-damaging activity in high concentrations from 40 mmol/l (Garberg et al. 1998).
In the alkaline elution assay EDTA disodium salt in a concentration of 30 mmol/l, with and without metabolic activation, showed no mutagenic activity (Swenberg et al. 1976).
EDTA disodium salt was studied in mice for mutagenic activity in a bone marrow micronucleus assay, a dominent lethal assay and in the incidence of spermhead abnormalities. The doses ((5-20 mg/kg body weight) were given orally. EDTA disodium salt induced a dose dependent increase in the incidence of micronucleated polychromatic erythrocytes, but no mutations in the dominant lethat assay and no increase in the spermhead abnormalities were seen (Muralidhara and Narasimhamurthy 1991).
Contrasting results are thus obtained concerning the mutagenicity of EDTA. Additional evaluation may be considered.
EDTA and its salt were studied for teratogenic potential in rats. The equimolar dose of 1,000 mg/kg body weight was given by gastric intubation twice daily on day 7 to 14 of gestation. No teratogenic effects occurred with any of the compounds even at maternally toxic doses (Shardein et al. 1981). Disodium EDTA was given to pregnant rats on day 7 to 14 of gestation by gavage (954 mg/kg body weight/day) and by subcutanous injection (375 mg/kg body weight/day). Disodium EDTA in the diet resulted in severe maternal toxicity and malformations in 71% of the offspring. Disodium EDTA given by gastric intubation (1,250 mg/kg/day or 1,500 mg/kg/day) was much more toxic to the dams. 87.5% maternal deaths but fewer malformed offspring. Disodium EDTA given subcutaneously was lethal to 24% of the dams at a much lower dose than given by either oral route, but did not produce a significant number of malformations in the offspring. For subcutaneous absorption the compound might not reach the embryo in concentrations high enough to produce malformations. A greater absorption of dietary EDTA into the circulation would correlate with the large increase in malformations following this route of administration. The route of administration had significant effect on the toxicity and teratogenicity (Kimmel 1977). It was suggested that the teratogenic effects of EDTA given to rats at very high levels were due to zinc deficiency. The binding of EDTA to zinc may be the most important interaction during pregnancy in that the developing embryo is extremely sensitive to zinc deficiency. Teratogenicity could be prevented by zinc diet supplement (Swenerton and Hurley 1971; Wolf and Gilbert 1992; WHO 1998).
EDTA salts are considered, by the Danish Labour Inspection Service, as a suspected reproductive toxicant at medium dose, meaning 20 to 200 mg/kg body weight (Arbejdsmiljøinstituttet 1990b).
EDTA is irritating to the eyes. The teratogenic potential of EDTA and its salts has been investigated but with variable results. EDTA and salts have been shown to be teratogenic after oral administration in rats. EDTA is not included in Annex 1 of list of dangerous substances of Council Directive 67/548/EEC. BASF classify EDTA, tetrasodium salt as Harmful (Xn) with the risk phrases R22 (Harmful if swallowed) and R36 (Irritating to the eyes) (BASF 1999). No data showed ability to induce sensitisation to human skin, but EDTA disodium salt is listed as a potential contact allergen by the Danish Labour Inspection Service (Arbejdsmiljøinstituttet 1990a).
Nitrilotriacetate, NTA, (CAS No. 139-13-9) is an organic compound belonging to the group of amino carboxylic acids, which have strong chelating capacity. Chelating agents react with polyvalent metal ions to form one or more ring structures. NTA acts by sequestering metal ions and is very effective in removing both calcium and magnesium from wash waters. In terms of washing performance NTA can largely replace phosphates (Perry 1981). However, NTA has received considerable attention primarily due to its demonstrated carcinogenicity and heavy metal chelating properties.
The strong complexing capacity of NTA is expected to have adverse effects upon heavy metal removal during sewage treatment and upon mobilisation of metals from sediments in receiving waters. Several investigations have shown that the presence of NTA in water/sediment systems increases the concentration of heavy metals in the water phase (Perry et al. 1984; Garnett et al. 1986; Dehnad and Radeke 1993). However, these experiments have usually been performed with a sediment water suspension. The fact that the mobilisation of metals from stable sediments into the water phase depends on the diffusion rate has not been taken into consideration. The diffusion from stable sediments is slow and several days are normally required before a steady state is achieved (Källqvist). An experiment with stable artificial sediment (kaolinite) indicated that high concentrations of NTA remobilized Zn and it was concluded that NTA concentrations above 200 m g/l might mobilise heavy metals from stable sediments (Bernhardt 1991, cited in Källqvist). A continuous exposure with NTA may enhance the risk of metal remobilization although the low diffusion of metals from the sediment reduces the transport of metals (Källqvist).
NTA is known to be aerobically biodegradable by acclimated microorganisms. Biodegradability tests with NTA have been inconsistent; 90% degradation has been reported after 9 and 13 days in tests with activated sludge, while degradation attained only 20% in a CO2 evolution test after 28 days and did not occur in shake flask and BOD tests (Perry et al. 1984). Following a period of acclimatisation, almost complete biodegradation has been reported for the activated sludge process when operated under optimum conditions. The efficiency of NTA removal during biological sewage treatment and the period of acclimatisation prior to NTA biodegradation has shown to be affected by factors like, e.g., the concentration of heavy metals, treatment temperature, NTA concentration and water hardness (Perry et al. 1984).
The removal of NTA during anaerobic sludge digestion has been found to be variable and affected by operational characteristics. E.g., studies indicating no removal and up to 29-45% removal in digesters receiving co-settled primary and activated sludge over a period of 120 days have been reported (Perry et al. 1984).
The toxicity of NTA towards algae, crustaceans and fish is low with EC/LC50 values well above 100 mg/l (Table 7.8).
Toxicokinetics and acute toxicity
Na3NTA is poorly absorbed from the gastrointestinal tract in humans. When absorbed the compound is rapidly excreted in the urine. About 87% of the absorbed dose were excreted within the first 24 h post dosing. NTA is not biotransformed and is excreted almost entirely unchanged in urine (Budny and Arnold 1973).
14C-labelled NTA was given intravenously and by stomach intubation to mice and the distribution was studied with autoradiography. Up to 48 hours after dosing a high concentration of radioactivity in the skeleton was seen. NTA has a preference for bone where it forms complexes with divalent cations such as calcium. In addition to the skeleton, a high concentration of radioactivity was seen in the kidney and the urinary bladder up to 8 hours after injection (Tjälve 1972).
The absorption, distribution and metabolic excretion of NTA in mice were determined by oral administration. Excretion of 14C-labelled NTA after a single oral dose showed that 99% of the dose was eliminated within 24 h. About 96% in the urine and the rest in faeces. NTA was readily absorbed from the gastrointestinal tract of the mice and was rapidly distributed into all tissues with highest concentrations in the bladder, kidney and bone. Elimination of NTA from the skeletal tissue was also rapid after 8 hours no detectable radioactivity was left. This indicates no serious accumulation in the bone (Chu et al. 1978). NTA is poorly absorbed in humans compared with experimental animals. The absorption through skin is minimal. Less than 0.1% of dermal doses are absorbed (Anderson and Alden 1989).
The acute toxicity of NTA and its salts in animals are relatively low. The acute toxicity (LD50 values) of NTA are given in Table 7.9.
NTA is a skin irritant. The degree depends on the neutralization (Richardson 1992-1994). A 20% solution of Na3NTA was not skin irritating in a patch test on 66 persons (Nixon 1971). NTA is a mild eye irritant (Grant and Schuman 1993).
Dermal exposure to NTA does not cause sensitization (Anderson and Alden 1989). A 20% solution of Na3NTA was not allergenic in a patch test on 66 persons (Nixon 1971).
Rats fed for 90 days with diets containing 2,000 ppm (0.2 g/kg bw/day) Na3NTA and no effects were observed. Rats fed a diet containing 20,000 ppm (2 g/kg bw/day) had abnormal kidneys and a significant decrease in weight gain with a corresponding increase in organ/body weight ratios (liver and kidney) (Nixon 1971).
Mutagenicity and Carcinogenicity
NTA induces tumours only after prolonged exposure to higher doses than those producing kidney toxicity. The reported induction of tumours in rodents is considered to be due to cytotoxicity resulting from the chelation of divalent cationics such as zinc and calcium in the urinary tract (WHO 1996). Dosages of NTA that do not alter Zn or Ca distribution do not produce any urinary tract toxicity even after chronic exposure. When toxic doses are supplied chronically some of the severely damaged tissues may develop tumours (Anderson et al. 1985). Rats were given 0.1% NTA trisodium salt in drinking water for 2 years. The exposed animals showed an increase in hyperplasia and tumourigenesis in the kidney (Goyer et al. 1981). Nitrilotriacetic acid and nitrilotriacetic acid, trisodium salt were tested for carcinogenicity in mice and rats by oral administration and induced tumours of the urinary system (kidney, ureter and bladder). The monohydrate administered in the diet induced malignant tumours of the urinary system. When administered in drinking water to rats, it induced renal adenomas and adenocarcinomas (IARC 1990).
The International Agency for Research on Cancer (IARC) has evaluated that there is sufficient evidence for the carcinogenicity of NTA and its sodium salts in experimental animals and the overall evaluation is that nitriloacetic acid and its salt are possibly carcinogenic to humans. IARC has placed NTA in Group 2B (IARC 1990).
The potential of NTA to cause chromosome abnormalities was investigated in cell culturs (human lymphocytes and Chinese hamster ovary cells) and in vivo in mice (micronucleus test). NTA was not found mutagenic in any of the three test assays (Monaldi et al. 1988; Loveday et al. 1989).
The effect on reproduction and development of Na3NTA in the diet was studied in rats for two generations and in rabbits during a single pregnancy. Na3NTA was fed to rats either continuously or only during organogenesis (from day 6 to 15) in each pregnancy at one or two dietary levels, 0.1 and 0.5%. For the rabbits doses of 2.5, 25, 100 and 250 mg Na3NTA/kg body weight were given by stomach tube during organogenesis (on day 7 to 16 of pregnancy). Na3NTA caused no effects on reproduction or embryonic development in either rats or rabbits. The only effects of Na3NTA on the rats were some growth depression in both adults and wealing animals fed 0.5% (Nolen et al. 1971). Pregnant mice were given 0.2% NTA in the drinking water from day 6 to 18 of pregnancy. The fetuses were examined for malformations. Skeletal or visceral examination did not reveal any teratogenic effects, although NTA also accumulated in the foetal skeleton (Tjälve 1972).
NTA was not found teratogenic in the frog embryo teratogenicity assay (Dawson et al. 1989).
Exposure to nitrilotriacetic acid, and presumably also to its water-soluble metal complexes, occurs as a result of its presence in household detergents and in drinking water. Little information on the toxicity of NTA in humans is available. The kidney is the primary target for NTA toxicity in animals. There is a clear evidence of carcinogenicity in rats and mice, causing kidney, bladder and urinary tract tumours in high doses and after long-term exposure. No human carcinogenic data are available. There is no evidence of teratogenicity and mutagenicity. The mechanism of the toxicity can be partly explained by chelation of essential divalent metal ions such as Ca++, Mg++ and Zn++.
Nitrilotriacetic acid with sodium salts is not included in Annex 1 of list of dangerous substances of Council Directive 67/548/EEC. Sodium salts of nitrilotriacetic acid are included in the list of carcinogenic components of the Executive Order on precautions to prevent cancer risk issued by the National Working Environment Authority (Executive Order 1999).
BASF classify NTA as Harmful (Xn) with the risk phrases R22 (Harmful if swallowed) and R36 (Irritating to the eyes) (BASF 1999).