Silver bromide

Endpoint summary

Additional information

Acute and chronic aquatic toxicity data were available for a range of different fresh and marine water species. Where possible, all toxicity test data used for deriving PNECs were from studies that used silver nitrate (AgNO₃) as the test material. Silver nitrate dissociates readily in aqueous solution to liberate Ag⁺ ions, the species of silver considered to be of greatest toxicity. Where data for silver nitrate were not available data derived from other salts (e.g. silver chloride) were used, but only after the exposure conditions were determined to be acceptable (e.g. no testing conducted beyond limits of solubility and Ag⁺ likely to be fully dissociated). All data are reported in terms of silver ion concentration and are read across to other less soluble silver compounds to assess their toxicity.

Where there were sufficient data, PNECs have been derived using statistical extrapolation (species sensitivity distributions, SSDs). In the absence of sufficient data to derive a PNEC using an SSD then a conventional deterministic (assessment factor) approach has been adopted. Both techniques are compliant with current ECHA guidance (ECHA, 2008). For full details of the calculated PNECs see attached appendix.

The effects database for marine species is considerably smaller than that for freshwater organisms. Long-term toxicity data are available for five different taxonomic groups: algae, crustaceans, echinoderms, fish and molluscs. The saltwater toxicity data suggest that silver toxicity is modified by salinity, with increased toxicity occurring at lower salinities (Ward et al., 2006b, Pedroso et al., 2007). Because of the speciation of silver in different media it is not recommended that the freshwater and saltwater datasets be combined when considering PNECs for saltwaters.

Potential Mitigating Effect of Sulphide on Chronic Silver Ecotoxicity

 

Some freshwater physico-chemical properties are thought to have an influence upon chronicAgtoxicity. As well as the usual properties affecting metal bioavailability in freshwaters, such as dissolved organic carbon, sulphide is also thought to have a potentially mitigating influence on the ecotoxcicity ofAg. Historically, one of the practical difficulties of assessing the influence of sulphide uponAgin freshwaters has been the complexity of the method of determination and complete absence of any high quality routine monitoring data. These problems have now been addressed by the National Laboratory Service of England and Wales and a study has been initiated collecting routine monitoring data (including sulphide measures) from a range ofAgexposures in freshwaters and using these data to empirically evaluating the potential for the developing a chronic biotic ligand model forAg.

Work is currently ongoing but initial results indicate that free sulphide (operationally defined as chromium reducible sulphide, CRS) could have a potentially significant mitigating effect upon the free silver ion concentration in solution. Further details are provided below.

 

The acute toxicity ofAghas been shown to be mitigated, to some degree, by elevated levels of sodium, chloride, DOC and, to a lesser degree, hardness. The limited effectiveness of hardness cations at mitigating acuteAgtoxicity is probably related to the high strength of binding ofAg⁺at the site of action, (i.e., to the biotic ligand), a characteristic that makes it difficult forAg⁺to be displaced by competing cations. Chromium reducible sulphide (CRS) is another water quality constituent that has the potential to be very important, particularly at relatively lowAgconcentrations, as a result of the high affinity of CRS forAg⁺.

 

The 2010 UK monitoring data (Simpson et al. 2010) have shown that the environmental levels of silver tend to be relatively low, in most cases less than the PNEC of 40 ng/L (0.04 µg/L, or ~0.37 nM). At 14.6 nmol/mg DOC (Kramer et al., 2007), and a typical range of DOC levels of 2 – 10 mg/L (or higher), it is expected that CRS levels will commonly be in the concentration range of about 30 – 150 nM, an estimate which is consistent with the monitoring results from theUKprogramme. The molar ratio of CRS toAg⁺should therefore be about 75 – 425 (using theAgPNEC as a conservatively high estimate of the ambientAg⁺concentration). Although other cationic metals will also bind to CRS, the analyses of the preceding section indicate that they are not effective at competing withAg⁺for CRS binding sites. Hence, even for the relatively high metal concentrations that were assumed above,Ag⁺should effectively compete with the other cationic metals for complexation by CRS and there should normally be enough CRS to complex theAg⁺that is present. This should in turn markedly reduce the concentration ofAg⁺, an important bioavailable form ofAg(Bianchini and Bowles, 2002). The question that remains to be addressed is whether or notAg-CRS complexes are sufficiently low in bioavailability and that the residualAg⁺concentration is sufficiently low such that an adequate level of protection againstAgtoxicity, chronic toxicity in particular, will be realized. Some of the experimental evidence that is relevant to consider in regard to this topic is summarized below.

 

Bianchini and coworkers investigated the protective effect of CRS on toxicity due toAg⁺. They considered the effect of CRS on the acute toxicity ofAg⁺toD. magna,a relatively sensitive invertebrate (Bianchini et al., 2002). When CRS was present at 25 nM it increased theAgLC50 by 5.6-fold relative to the LC50 in water without CRS (from 0.16 and 0.26 µg/L to 1.47 µg/L). This increase is comparable to but somewhat less than theAg⁺complexation capacity of the CRS (2.7 µg/L). The difference may in part reflect uncertainty in the measured concentrations, given that changes inAgconcentration have been shown to occur over the course of toxicity tests (Bowles et al., 2002). It may also reflect the allowance of an insufficient time (3 hours) for pre-equilibration of theAg-CRS complex prior to daily water renewal. This latter explanation is consistent with the results of Glover et al. (2005) who showed an increase in LC50 of approximately 55% to 100% (as much as a factor of two increase, depending on DOC level) when the equilibration time betweenAgand NOM (presumably associated with CRS) was increased from 3 hours to 24 hours.

 

Importantly, when the CRS concentration in the preceding study exceeded the totalAgby more than 10-fold (250 nM vs 18.5 nM), it provided complete protection toD. magnaover the full range of the silver concentrations used to define the dose-response curve in the absence of CRS (i.e., over a range of 0 - 2 µg/LAg). This provided a clear demonstration of the protective effect of CRS, at least with regard to acute toxicity ofAgto a relatively sensitive invertebrate. Interestingly,Agwas accumulated by the D. magna to a greater degree in the presence of CRS, a result attributed to sulphide-bound silver in the digestive tract of the daphnids rather thanAgbound to the exoskeleton (Bianchini et al., 2004). The observation thatAgaccumulation occurred to a greater extent in the presence of elevated CRS (when toxicity was reduced) is an indication that the accumulation occurred via a physical process such as ingestion of colloidalAg, and that the accumulatedAgwas not toxicologically available to the organism. Additional measurements demonstrated that the gastrointestinal tractAgburden was in fact markedly elevated while depuration experiments showed that it was readily eliminated from the organism over a time course of several hours (Bianchini et al., 2004).

 

With regard to chronic toxicity, a more recent study evaluated the protective effects of either hardness or CRS (at an environmentally relevant concentration of 23 nM, as might be associated with DOC ~ 1.6 mg/L) on the acute (48-hour) and chronic (21-day) toxicity ofAgtoD. magna(Bianchini and Wood, 2008). Effect levels were computed on the basis of both total and dissolvedAg. On the basis of a 1:1 stoichiometry forAgHS (Bowles et al., 2002a) the amount of CRS added to the test water was equivalent to 2.48 µg/LAg⁺complexation capacity. TheD. magnaeffect levels for dissolvedAgin the presence of CRS were essentially the same as forAgin the absence of CRS due to the demonstrated sorption ofAgCRS to membrane filters and other surfaces (Bowles et al., 2002a,b). As a result, the dissolvedAgdata are not of direct use in an assessment of the level of protection provided by CRS. On a totalAgbasis, however, the effect level for the 48-hour EC50 for survival was increased from 6.9 to 8.3 µg/L in the presence of CRS. This increase of 1.4 µg/L is somewhat less than the estimated complexation capacity.Again, the difference may have resulted from the relatively shortAg-CRS equilibration time that was provided (3 hours) when the exposure water was renewed each day, as this would limit the degree ofAgcomplexation that could occur. The 21-day chronic tests exhibited a result that was more consistent with stoichiometric expectations, with the 21-day survival EC50 increased by 2.23 µg/L in the presence of CRS (slightly less than the estimated complexation capacity of CRS).

 

The 21-day chronic toxicity test results forD. magnaalso showed that that CRS provided a protective benefit for several of the end points that were reported. The tests were performed with and without the addition of 23 nM of CRS (complexation capacity ~ 2.48 µg/L). End points included total number of neonates produced (TN), time to first brood (TB), number of broods (NB), number of young per brood (YB) number of reproduction days (RD) and number of young per adult per reproduction day (YAD). The differences in totalAgLC50s (with and without CRS) in 48-hour and 21-day tests were 1.4 µg/L and 2.23 µg/L, respectively. The increase for sub-lethal effects on reproduction (YAD) was 1.65 µg/L. The collective average of 1.8 µg/L is similar to the complexation capacity of 23 nM CRS (2.48 µg/L).Again, the somewhat lower value that was observed likely reflects the short equilibration time that was provided when the test water was renewed each day. Other end points exhibit a limited benefit of the added CRS. However, theAgeffect levels were relatively high in comparison to the CRS levels, so this is not surprising.

 

Chronic toxicity study results have also been reported forC. dubia(7-day static renwal) exposed toAgin the presence and absence of sulphide (added as CuS; Naddy et al., 2007). The CRS increase from <2.2 nM (Horsetooth Reservoir control water) to 75.4 nM (8.14 µg/L complexation capacity) led to a 43% increase in the IC20. The NOEC was 17.5 µg/L. The CRS complexation capacity of 8.14 µg/L was slightly higher than the increase in the survival and reproduction NOECs of 6.1 and 5.8 µg/L, respectively. Once again, only a 3 hour equilibration time was used, so the full benefit of the added CRS may not have been realized. Even so, the results further demonstrate the protective benefit of CRS.

 

With regard to theUKmonitoring data, ambient silver levels are in most cases less than or about equal to the PNEC of 40 ng/L (0.04 µg/L). Further, CRS is typically present at much higher concentrations than theAg, such that CRS is available in considerable excess. As was seen above for the acute toxicity results, when CRS was present at a 10-fold higher concentration than the effect level,Agtoxicity was essentially eliminated. Based on these results, toxicity due to silver is likely to be effectively limited by CRS in most settings. When considered that other mitigating factors may also be present, chronic toxicity due to silver should be reduced significantly.

 

Reference Cited

Bianchini, A., and K.C. Bowles, 2002. “Metal Sulfides in Oxygenated Aquatic Systems: Implication for the Biotic Ligand Model,”Comparative Biochemistry and Physiology, Part C 133: 51-64.

Bianchini, A., K.C. Bowles, C.J. Brauner, J.W. Gorsuch, J.R. Kramer and C.M. Wood, 2002. “Evaluation of the Effect of Reactive Sulfide on the Acute Toxicity of Silver (I) toDaphnia magna.Part 2: Toxicity Results,”Environmental Toxicology and Chemistry, 21(6): 1294-1300.

Bianchini, A., C. Rouleau and C.M. Wood, 2004. “Silver Accumulation inDaphnia magnain the Presence of Reactive Sulfide,”Aquatic Toxicology, 72(4): 339-349.

Bianchini, A., and C. M. Wood. 2008. “Does sulfide or water hardness protect against chronic silver toxicity inDaphnia magna? A critical assessment of the acute-to-chronic toxicity ratio for silver,”Ecotoxicology and Environmental Safety71:32-40.  

Bowles K.C., A. Bianchini, C.J. Brauner, J.R. Kramer, C. M. Wood. 2002a. “Evaluation of the effect of reactive sulfide on the acute toxicity of silver (I) toDaphnia magna. Part 1: description of the chemical system,” Environmental Toxicology and Chemistry21:1286-1293. 

Bowles, K.C., R.A. Bell, M.J. Ernste, J.R. Kramer, H. Manolopoulos andN. Ogden, 2002b. “Synthesis and Characterization of Metal Sulfide Clusters for Toxicological Studies,”Environmental Toxicology and Chemistry, 21(4): 693-699.

Glover, C., R.C. Playle and C.M. Wood, 2005. “Heterogeneity of Natural Organic Matter Amelioration of Silver Toxicity toDaphnia magna: Effect of Source and Equilibration Time,”Environmental Toxicology and Chemistry, 24(11): 2934-2940.

Kramer J.R., R.A. Bell, D.S. Smith. 2007. “Determination of sulfide ligands and association with natural organic matter,” Applied Geochemistry 22:1606-1611.

Naddy, R.B., J.W. Gorsuch, A.B. Rehner, G.R. McNerney, R.A. Bell and J.R. Kramer, 2007. “Chronic Toxicity of Silver Nitrate toCeriodaphnia dubiaandDaphnia magna, and Potential Mitigating Factors,”Aquatic Toxicology, 84: 1-10.

Simpson, P., Brown, B., Peters, A. and Merrington, G. 2010. Silver emissions to freshwaters inEnglandandWales. A report to the EnvironmentAgency of England andWalesand the Precious Metals Consortia from WCA Environment Ltd. Report Number P0181_09-10.