Toxicant default guideline values for sediment quality
Default guideline values (DGVs) for toxicants in sediment are listed in Table 1, together with their origins. As per the DGVs for toxicants in water, the sediment DGVs indicate the concentrations below which there is a low risk of unacceptable effects occurring, and should be used, with other lines of evidence, to protect aquatic ecosystems. In contrast, the ‘upper’ guideline values (GV-high), also listed in Table 1, provide an indication of concentrations at which you might already expect to observe toxicity-related adverse effects. As such, the GV-high value should only be used as an indicator of potential high-level toxicity problems, not as a guideline value to ensure protection of ecosystems.
We recommend using multiple lines of evidence as part of the weight-of-evidence process to better assess the risk to a sediment ecosystem if a DGV is exceeded or even where toxicant concentrations in the sediment are trending towards the DGV.
How we derived these default guideline values
ANZECC & ARMCANZ (2000) introduced toxicant DGVs for sediments but only a limited number of sediment toxicity tests were available at that time. It was not possible to derive reliable guideline values using species sensitivity distributions (SSDs) of chronic toxicity data, as completed for toxicants in water.
Instead, ANZECC & ARMCANZ (2000) derived DGVs using a ranking of both field ecological and laboratory ecotoxicity-effects data from North America (Long et al. 1995), with the DGV representing the 10th percentile value of the data distribution and using the median value as an additional upper guideline value (GV-high).
Variants to this approach are being used in many jurisdictions internationally. Numerous studies have demonstrated that below the DGV, effects on sediment biota are rarely seen, while effects are more frequently evident above the GV-high value.
When Simpson et al. (2013a) revised the 2000 sediment quality guidelines, they detailed some revisions to the DGVs and their application in sediment quality assessment.
Since 2000, it became clear that high variability existed internationally between the various guideline value derivations for organic toxicants in sediments. The effects range low (ERL) values of Long et al. (1995) are considered to be less reliable than the threshold effects level (TEL) values of MacDonald et al. (2000) that were adopted in Canada (CCME 2002).
In the Water Quality Guidelines, we have:
- adopted the values from McDonald et al. (2000) for many organic toxicants, as in Simpson et al. (2013a)
- revised the DGVs for polycyclic aromatic hydrocarbons (PAHs) and tributyltin (TBT)
- reported a new value for total petroleum hydrocarbons (TPHs).
With the availability of more sediment toxicity tests, the use of SSDs to derive guideline values has been demonstrated for copper (Simpson et al. 2011). At the same time, the role of grain size and organic carbon (OC) as the major variables controlling contaminant bioavailability was illustrated as a precursor to future guideline value derivations that might more explicitly take these into account.
Applying the sediment quality guideline values and adjusting for different levels of protection
The toxicant concentrations measured from the <2 mm sediment fraction of a sediment sample should be compared with the sediment DGVs.
As the DGVs for sediments have not been derived using SSDs, it is not possible to adjust the level of protection based on the percentage of species protected. However, the level of protection can be adjusted if sediment quality guideline values are based on reference site data (see below).
The GV-high represents the median value of the effects ranking. As such, GV-high could be considered as more likely to be associated with biological effects than the DGV but the extent of that impact is not necessarily known. You should only use the GV-high as an indicator of potential high-level toxicity problems, not as a guideline value to ensure protection of ecosystems. We recommend a multiple lines-of-evidence approach to better assess the risk to a sediment ecosystem if a DGV is exceeded
Refining guideline values for local conditions
Sediment grain size
Bioavailability and toxicity of contaminants is influenced by sediment grain size. The contaminant binding capacity of sediments decreases with increasing grain size, and this results in the concentration of contaminants typically being greater in the finer sediment fractions. An exception to this is where the contamination source is particulate, for example, paint particles and industrial solids.
The data used in the effects database were largely associated with silty rather than sandy sediments so the DGVs are most applicable to silty sediments.
In some cases, more stringent guideline values might be applied to sandy sediments because partitioning to pore waters (water contained in pores in soil or rock) is more favourable.
The <2 mm sediment particle size fraction should be used for chemical analyses for comparison with sediment quality guideline values so that the potential risk posed by contaminants is not diluted by a large mass of larger materials (gravel and other debris). The <63 μm sediment particle size fraction (clay and silt) is considered a suitable representation of the sediment materials that are mostly readily resuspended or potentially ingested by organisms. Therefore, it is recommended that the binding of contaminants by this fine sediment fraction be considered when more detailed investigations of contaminant bioavailability are required for site-specific assessments (see Simpson et al. 2013a for more detailed guidance).
Organic carbon content
Increasing OC content favours partitioning of both metals and organics to sediment particles.
For hydrophobic organic contaminants (HOCs), the DGVs are normalised to 1.0% OC (Simpson & Batley 2016). Therefore, the measured HOC concentration in the sediment also needs to be normalised to the OC content.
For sediments with 0.2 to 10.0% OC (dry weight), normalising the measured HOC concentration to 1% OC is undertaken by dividing the measured HOC concentration by the measured percent OC. If the sediment OC concentration is below or above this range, the measured HOC concentration should be normalised to (i.e. divided by) 0.2% and 10.0 %, respectively.
The OC content range of 0.2 to 10.0% is judged to be necessary because second-order effects (for example particle size and adsorption to non-organic mineral fractions) become more important at lower OC contents, while the validity of the approach weakens above 10% OC.
You can calculate normalisation by dividing the HOC concentration by the percentage of OC content for cases within the described limits.
Examples:
If sediment contains 5 mg/kg of total PAHs and 0.55% OC, then 1% OC normalised concentration = 5/0.55 = 9.1 mg/kg of total PAHs (1% OC).
If sediment contains 15 mg/kg of total PAHs and 3.2% OC, then 1% OC normalised concentration = 15/3.2 = 4.7 mg/kg of total PAHs (1% OC).
Simpson et al. (2013b) derived guideline values for copper in sediment expressed as mg < 63 µm Cu/kg OC but found that the existing DGV was adequately protective. Therefore, this approach of normalising to OC has not been used for copper or any other metal contaminant in Table 1.
Absence of toxicant values for sediments
In some cases, no DGVs will be specified for a toxicant of interest. This generally reflects absence of an adequate dataset for that toxicant.
An interim approach is required to provide some guidance and ensure environmental protection in situations where guideline values would apply.
The suggested approach is to use the reference site approach, where a site-specific guideline value is derived on the basis of natural background (reference) concentration. We recommend using the 80th percentile of the reference site concentration (which is then compared with the median concentration at the site of interest). Such guideline values need to be based on an adequate number of samples to ensure statistical robustness (see associated guidance for sediments in the Data analysis section of Monitoring). Depending on the agreed ecosystem condition and associated level of protection, this percentile can be adjusted down or up to afford higher or lower protection, respectively. However, decisions on the appropriate percentile should be consistent with the guidance for deriving guideline values using reference site data.
It is important to compare your derived guideline values with concentrations for sediments of similar grain sizes and OC contents.
An alternative approach is to apply the water quality DGVs to sediment pore waters. This is based on equilibrium partitioning theory (Di Toro et al. 1991), which assumes that the critical factor controlling sediment toxicity is the concentration of toxicant in the pore water. This method can involve:
- measuring toxicant concentrations in pore waters, or
- using a known (or measured) water sediment partition coefficient to calculate the pore water concentration from the measured pore water concentration.
Recommended default guideline values for toxicants in sediment
| Type of toxicant | Toxicant | DGV | GV-high |
|---|---|---|---|
| Metals (mg/kg dry weight)a | Antimony | 2.0 | 25 |
| Cadmium | 1.5 | 10 | |
| Chromium | 80 | 370 | |
| Copper | 65 | 270 | |
| Lead | 50 | 220 | |
| Mercury (Inorganic) | 0.15 | 1.0 | |
| Nickel | 21 | 52 | |
| Silver | 1.0 | 4.0 | |
| Zinc | 200 | 410 | |
| Metalloids (mg/kg dry weight)a | Arsenic | 20 | 70 |
| Organometallics (µg/kg dry weight, 1% OC)c, d | Tributyltin (as Tin) | 9.0 | 70 |
| Organics (µg/kg dry weight, 1% OC)b,c | Total PAHse | 10,000 | 50,000 |
| Total DDT | 1.2 | 5.0 | |
| p.p’-DDE | 1.4 | 7.0 | |
| o,p’- + p,p’-DDD | 3.5 | 9.0 | |
| Chlordane | 4.5 | 9.0 | |
| Dieldrinf | 2.8 | 7.0 | |
| Endrinf | 2.7 | 60 | |
| Lindane | 0.9 | 1.4 | |
| Total PCBs | 34 | 280 | |
| Organics (mg/kg dry weight) b | TPHsg | 280 | 550 |
DDD = dichlorodiphenyldichloroethane; DDT = Dichlorodiphenyltrichloroethane; DDE = dichlorodiphenyldichloroethylene; DGV = default guideline value; GV-high = additional upper guideline value; PAHs = polycyclic aromatic hydrocarbons; PCBs = polychlorinated biphenyls; TPHs = total petroleum hydrocarbons; OC = organic carbon
a. Primarily adapted from the effects range low (ERL) and effects range median (ERM) values of Long et al. (1995).
b. Primarily adapted from threshold effects level (TEL) and probable effects level (PEL) values of MacDonald et al. (2000) and CCME (2002).
c. Normalised to 1% OC within the limits of 0.2 to 10%. Thus if a sediment has (i) 2% OC, the ‘1% normalised’ concentration would be the measured concentration divided by 2, (ii) 0.5% OC, then the 1% normalised value is the measured value divided by 0.5, (iii) 0.15% OC, then the 1% normalised value is the measured value divided by the lower limit of 0.2.
d. Basis of revision is described in Appendix A2 of Simpson et al. ( 2013a).
e. The DGV and GV-high values for total PAHs (sum of PAHs) include the 18 parent PAHs: naphthalene, acenaphthylene, acenaphthene, fluorene, anthracene, phenanthrene, fluoranthene, pyrene, benz[a]anthracene, chrysene, benzo[a]pyrene, perylene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[e]pyrene, benzo[ghi]perylene, dibenz[a,h]anthracene and indeno[1,2,3-cd]pyrene. Where nonionic OCs like PAHs are the dominant chemicals of potential concern (COPCs), the use of equilibrium partitioning sediment benchmarks (ESBs) is desirable, which includes a further 16 alkylated PAHs (generally listed as C1-/C2-/C3-/C4-alkylated ), as described in Appendix A3 of Simpson et al. (2013a).
f. Where dieldrin or endrin are the major COPCs, it is recommended that ESB approaches are applied as described in Appendix A4 of Simpson et al. ( 2013a).
g. Origin described in Appendix A5 of Simpson et al. ( 2013a).
References
ANZECC & ARMCANZ 2000, Australian and New Zealand Guidelines for Fresh and Marine Water Quality, Australian and New Zealand Environment and Conservation Council and Agriculture and Resource Management Council of Australia and New Zealand, Canberra.
CCME 2002, Canadian Sediment Quality Guidelines for the Protection of Aquatic Life — Summary Tables, Update 2002, Canadian Council of Ministers of the Environment, Winnipeg, Canada.
Di Toro, DM, Zarba, CS, Hansen, DJ, Berry, WJ, Swartz, RC, Cowan, CE, Pavlou, SP, Allen, HE, Thomas, NA & Paquin, PR 1991, Technical basis for establishing sediment quality for non-ionic organic chemicals using equilibrium partitioning, Environmental Toxicology and Chemistry 10: 1541−1583.
Long, ER, MacDonald, DD, Smith, SL & Calder, FD 1995, Incidence of adverse effects within ranges of chemical concentrations in marine and estuarine sediments, Environmental Management 19: 81–97.
MacDonald, DD, Ingersoll, CG & Berger, TA 2000, Development and evaluation of consensus‐based sediment quality guidelines for freshwater ecosystems, Archives of Environmental Contamination and Toxicology 39: 20–31.
Simpson, SL & Batley, GE 2016, Sediment quality assessment; a practical handbook, CSIRO Publishing, Clayton, Vic., 346 pp.
Simpson, SL, Batley, GE, Hamilton, I & Spadaro, DA 2011, Guidelines for copper in sediments with varying properties, Chemosphere 85(9): 1487–1495.
Simpson, SL, Batley, GE & Chariton, AA 2013a, Revision of the ANZECC/ARMCANZ Sediment Quality Guidelines, CSIRO Land and Water Report 8/07, CSIRO Land and Water.
Simpson, SL, Spadaro, DA & O’Brien, D 2013b. Incorporating bioavailability into management limits for copper in sediments contaminated by antifouling paint used in aquaculture, Chemosphere 93(10): 2499–2506.
