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 to help ensure the protection of 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.

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. (2013)
  • 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.

Adjusting for different levels of protection

Unless DGVs for sediments have been derived using SSDs applied to an adequate toxicity dataset it is not possible to derive guideline values for different levels of species protection. (This differs from the toxicant guideline values for water quality.)

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 depends primarily on sediment grain size. 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.

Sediment binding capacity decreases with increasing grain size. So you should apply more stringent guideline values to sandy sediments because partitioning to pore waters (water contained in pores in soil or rock) is more favourable.

The concentration of contaminants is typically much greater in the finer sediment fractions. Consequently, the >2 mm sediment fraction that may contain coarse sand and large debris is excluded from the chemical analyses. The <63 μm sediment fraction is considered a suitable representation of the sediment materials that are mostly readily resuspended or potentially ingested by organisms. It is recommended that the binding of contaminants by the fine sediment fraction be considered when more detailed investigations of contaminant bioavailability are required for site‐specific assessments.

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 (Batley & Simpson 2016).

For sediments with 0.2 to 10.0% OC (dry weight), normalising the DGVs to 1% OC has been shown to be valid for a range of sediment types. (Outside this range, the values 0.2% and 10.0% are used for normalisation).

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.

You can calculate normalisation by dividing the HOC concentration by the percentage of OC content for cases within the described limits.


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. This approach 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 guidelines would apply.

Our suggested approach is to derive a site-specific guideline value on the basis of natural background (reference) concentration multiplied by an appropriate factor. We recommend using a multiplying factor of 2. In some highly disturbed ecosystems, a slightly larger factor may be more appropriate (but no larger than 3).

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.

Revised default guideline values for toxicants in sediment

Table 1 Revised toxicant default guideline values for sediment quality
Type of toxicant Toxicant DGV GV-high
Metals (mg/kg dry weight)a Antimony 2.0 25.0
Cadmium 1.5 10.0
Chromium 80 370
Copper 65 270
Lead 50 220
Mercury 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 Tributyltin (µg Sn/kg dry weight, 1% TOC)c, d 9.0 70
Organics (µg/kg dry weight, 1% TOC)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
TPHs (mg/kg dry weight)g 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; TBHs = total petroleum hydrocarbons; TOC = total organic contaminants
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. (2013).
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).
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. (2013).
g. Origin described in Appendix A5 of Simpson et al. (2013).


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, 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.