weight of evidence analysis avs normalization nickel 28-06

WWW.ARCHE-CONSULTING.BE

WEIGHT OF EVIDENCE ANALYSIS OF THE SEDIMENT

BIOAVAILABILITY NORMALIZATION APPROACH USING THE

SEM-AVS MODEL FOR NICKEL IN SEDIMENTS

28th June 2012

Marnix Vangheluwe

[email protected]

COMMISSIONED BY

NIPERA

WEIGHT OF EVIDENCE ANALYSIS SEM-AVS CONCEPT FOR NICKEL

2

1. EXECUTIVE SUMMARY

The SEM-AVS concept was developed to predict situations in which toxicity should not occur. Metals

in sediment will not be toxic if the molar concentration of AVS is higher than that of SEM

(SEM/AVS ratio smaller than 1) or if the difference between the molar concentrations of SEM and

AVS (SEM-AVS) is used (Hansen, 1996) the molar SEM-AVS difference < 0. It should be

noted from the start that the SEM-AVS model was not designed to predict toxicity thresholds nor

was it developed to predict absence of bioaccumulation.

In the present report, a review of the use of the SEM/AVS concept for nickel is given. Literature

data are available describing the SEM/AVS concept for nickel, including acute and chronic studies in

both freshwater and marine sediments, as well as nickel specific field recolonization studies. An in-

depth evaluation of these literature data was performed in order to determine whether the existing

information validates the applicability of the SEM/AVS concept for nickel.

The recent body of work shows how the binding strength of reactive acid volatile sulfides (AVS)

can control pore water concentrations of divalent metals such as nickel. In these studies, AVS

concentrations consistently limit metal toxicity to invertebrates in anoxic and suboxic sediments where

molar AVS concentrations exceed molar concentrations of extractable metals. Although the SEM –AVS

models has been developed solely on establishing relationships between toxicity and dissolved metals

in pore water the preponderance of studies do also indicate reduced accumulation of metals at

sediment metal/AVS ratios of less than 1 (Ankley et al, 1996b). There are exceptions to this

general observation (Pesch et al, 1995, Lee et al 2000a, 2000b, De Jonge et al 2009,2010 and

2011)which are not surprisingly since steady state tissue concentrations are a function of uptake from

dissolved and dietborne exposures. Indeed the dietary route seems to play an important role in

explaining the observations that metals are taken up from the sediment irrespective of the AVS

concentration. It should, however, kept in mind that bioaccumulation does not represent a toxicological

effect and an unambiguous connection between observed levels of accumulation and effects has not

yet been made. At SEM-AVS concentrations < 0 metals extracted in the gut from the ingested

metal sulfides are detoxified and stored in granules while at an overload of the AVS system metals

can be found in a more toxicologically relevant pool. Overall the recent results seem to support the

tenet that the SEM/AVS approach can be used to separate situations where Ni toxicity can occur,

and that the approach can be used in a risk assessment framework.


3


4

2. INTRODUCTION

Assessing risks from metals for the sediment compartment are quite often hampered by the fact that

no clear relationship has been established between measured total concentrations of metals in

sediments and their potential to provoke toxic effects on aquatic life. As a result comparing

environmental concentrations expressed on a dry or wet weight basis with an established safety level

has the potential to result in an under or overestimation of the associated risk. It has been

recognized that for a realistic risk assessment for metals currently used procedures, solely based on

total concentrations, have to be improved by taking into account the bioavailable fraction of the

metals present in the sediments.

Attempts to address the inherent fundamental deficiencies of dry- or wet weight assessment

approaches are the use of normalization procedures based on one of the factors of the bulk

sediment matrix controlling the bioavailability of the chemicals of concern. Di Toro et al. (1991)

identified organic carbon as a key element controlling the bioavailability of non-ionic organic chemicals

and used this understanding in order to establish Sediment Quality Criteria for organic substances

using the equilibrium partitioning theory. A similar approach can be applied to metal contaminated

sediments where the sequestration of metals with Acid Volatile Sulfides (AVS) have been

demonstrated as being one of the important factors besides organic carbon controlling metal toxicity.

The underlying principle is that except for pyrite, naturally occurring iron and manganese mono

sulfides have higher solubility products than other metals (e.g. cadmium, copper, nickel, lead, zinc)

and can be displaced by these metals on a mole-to-mole basis, forming insoluble sulfide complexes

with minimized biological availability (Di Toro et al ,1990 & 1992 Ankley et al, 1996).

The applicability/usability of the SEM-AVS concept within a risk assessment context for metals has

been questioned consistently over the past 20 years. Some of the criticisms that have been

conveyed include:

• The dynamic nature of AVS: AVS concentrations have shown temporal and spatial

(horizontal and vertical) variations depending on sediment type and hydrological conditions

Van Den Berg et al, 1998 and Van Den Berg et al, 2001, Howard & Evans, 1993; Van

Den Hoop et al, 1997; Grabowski et al, 2001).


5

• Some sediment organisms create a micro-oxic environment by bioturbation (Peterson et al,

1995). This micro-habitat maximizes the exchange with the overlying water and hence

minimizes the mitigating capacity of a bioavailability factor such as AVS.

• It is not possible to preclude unambiguously other routes of uptake. AVS bound metals

may still be available to organisms that ingest sediment particles as evidenced by the

occurrence of bioaccumulation even when the sulfide pool exceeds the metals pool(Griscom

et al, 2000 & 2002, De Jonge 2009, 2010 and 2011)

The spatial and temporal variability is further discussed in Annex A. The next chapters explain in

more detail the mechanisms and processes underpinning the SEM-AVS concept and presents the

available nickel specific information with regard to the mitigating effect of AVS on nickel toxicity. The

main subject, however, of the current review is to elucidate the mechanisms of uptake and

bioaccumulation of metals and nickel in particular under conditions that SEM-AVS <0 and its

consequences to use the SEM-AVS model in a risk assessment context. SEM-AVS conceptual

framework.

2.1. DEFINITIONS

In anoxic sediments, sulfide produced by sulfate reduction reacts with Fe2+ and Mn2+ to form iron

and manganese sulfide solids, including amorphous iron sulfide, mackinawite, greigite, pyrrthotite,

troilite, pyrite and pink and green manganese sulfides (Wang and Chapman, 1999). Although pyritic

sulfide phases are both abundant and reactive towards trace metals, iron monosulfides, quite often

referred to with the term Acid Volatile Sulfides, are considered to be the more reactive sulfide pool.

According to Huerta-Diaz et al. (1998) AVS equals to the summation of amorphous FeS,

mackinawite (FeS) and greigite (Fe3S4). Acid volatile sulfide is, however, in the first place an

operational defined parameter indicating those sulfides that are readily extracted by the cold extraction

of sediment in approximately 1 M HCl acid..

Another term that is used in conjunction with AVS is SEM. SEM (Simultaneously Extracted Metal)

can be defined as the metal, which is simultaneously extracted under the same conditions under

which the AVS content is determined. If multiple metals are present it is necessary to use the term

total SEM (Σ SEM). The equivalent release of sulfide (AVS) and metal, however, does not

necessarily means that the metal is bound by sulfide alone. SEM refers to the metal associated with

the sulfides and any other metal-bearing phase that is extracted in the cold HCl extraction used for


6

AVS analysis (Allen et al., 1993). For example, metals adsorbed to iron oxides and particulate

organic carbon will also be extracted.

In general the release of nickel from its nickel sulfide complex is very limited. Simpson (2001)

mention recoveries of 1 ± 0.1 % (as percentage of the total concentration measured in sediment

digests n = 35 sediments) for his direct AVS screening method which is similar to the purge-and-

trap AVS procedure (1 M HCl) commonly used. For copper similar low recoveries has been

reported (0.8 ± 0.2%). Buykx et al (1996) expressed recovery rates as percentage of total

nickel concentration after sediment digestion. These results shown that SEMNi represents 4.4-31.1 %

of total nickel (n = 3 sediments). But in this case 6 M HCL has been used. For a risk

assessment purpose it can be assumed that the not soluble Ni in the HCl, extract will not be

toxicologically relevant.

2.2. OVERALL SEM-AVS CONCEPTUAL FRAMEWORK

Di Toro et al. (1990, 1992) have proposed the SEM/AVS model based on the recognition that

AVS is a reactive pool of solid phase sulfide available to bind with metals and hence reduce free

metal ion concentrations. The underlying principle is that except for pyrite, all other iron and

manganese mono sulfides have higher solubility products and can be displaced by other metals (e.g.

cadmium, copper, nickel, lead, zinc) on a mole-to-mole basis, forming insoluble sulfide complexes

with minimal biological availability (Ankley et al., 1996):

2/nMen+ + FeS (s) = Me2/nS(s) + Fe2+

2/nMen+ + MnS (s) = Me2/nS(s) + Mn2+

If all the metal in sediment is in the form of Me2/nS(s) (i.e. AVS in excess), then the free

metal ion activity is controlled by dissolution of Me2/nS(s).

The SEM-AVS concept was developed to predict situations in which toxicity should not occur. Metals

in sediment will not be toxic if the molar concentration of AVS is higher than that of SEM


7

(SEM/AVS ratio smaller than 1) or if the difference between the molar concentrations of SEM and

AVS (SEM-AVS) is used (Hansen, 1996) the molar SEM-AVS difference < 0.

Literature is replete with examples where it has been demonstrated that the SEM-AVS method is

effective in predicting the absence of metal toxicity in sediments. Mainly acute toxicity data were

used to establish the SEM-AVS model framework (Di Toro et al., 1990; Di Toro et al, 1991;

Casas and Crecelius, 1994; Pesch, et al., 1995; Berry et al., 1996, Ankley et al., 1996, Lee et

al, 2005). It should be stressed that the underlying basis for the model is solely based on

relationships between toxicity and predicted pore water concentrations. In the acute studies high metal

spikes were used that could have put more emphasis on the pore water route but also chronic and

fields tests support the use of the SEM-AVS model as a predictive tool for sediments that are

unlikely to be toxic (Vandegehuchte et al, 2007, Besser, 2011, De Witt et al, 1996). Field based

experiments using sediment spiked with individual metals and metal mixtures have also been

conducted and further validate the use of the SEM-AVS concept (Hare et al., 1994; De Witt et

al., 1996; Sibley et al., 1996; Hansen et al., 1996; Liber et al., 1996, Boothman et al., 2001,

Burton et al, 2005). For nickel, specifically, a large body of evidence has recently become available

that the SEM-AVS concept can be applied to nickel contaminated sediments (Besser et al , 2011;

Costelllo et al , 2011 and Nguyen et al, 2011). These studies are discussed in more detail in

chapter 4.

Within all these experimental designs other binding phases were intentionally excluded in the original

model as this would have limited the predictive capacity of the models.

As shown in Figure 1 The SEM/AVS concept does reasonably well in predicting the absence of

toxicity relative to total metal concentrations where no predictive ability is shown for toxicity in

sediments. Figure 1a (at the top) shows no relation at all between the total metal concentration

and the observed toxic effects. Normalizing for the AVS concentrations present results in a collapse

of all concentration response curves into one concentration response curve (Figure 1b) showing

absence of toxicity under SEM-AVS conditions < 0. Testing of 125 different field-sediments showed

the AVS approach accurately (99.2%) predicted the absence of toxicity (Hansen et al, 1996).


8

Figure 1 a and b: a) concentration response curves non normalized to AVS b)Effect of AVS normalization

on the predictive capacity of sediment toxicity

In a few separate cases toxicity was observed in field experiments while according to the SEM-AVS

model toxicity due to porewater metal exposure should not have occurred. For example in Liber et al

(1996) in 2 out of 17 freshwater field sites Zn toxicity was observed when excess AVS was

present. The endpoint was abundance of an oligochaete, but the response was not consistent through

the sampling events and the SEM/AVS ratios were very close to 1. This does not automatically

imply that the SEM/AVS model is flawed. One possibility is that the observed toxicity is not caused

by metals but is due to the presence of other contaminants that have not been measured. Or it

could also be partly due to sampling resolution in which the mean measured SEM-AVS values may

not always reflect what a benthic organisms actually 'see'. The bulk of evidence, however, in the

field sediments is supportive of the SEM/AVS model.

The applicability of the SEM/AVS model has also been evaluated by Shine et al (2003). In this

paper Receiver Operating Characteristics Curves (ROC) curves were used to compare different

approaches/models that estimate the toxicity of metals in sediments. The focus of the evaluation was

DRY WEIGHT NOR MALIZATIO N

AC ID VO LATIL E SULFIDE NOR MALIZATIO N

100

100

80

80

60

60

40

40

20

20

0

0

0.1

0.01

1

0.1

10

1

100

10

1000

100

C d

C d

C u

C u

Z n

Z n

N l

N l

C d + N l

C d + N l

Org

anis

m M

ort

ality

(%

)O

rgan

ism

Mor

talit

y (%

)

Sed ime nt ( mol/g)µ

S ed im ent ( m ol SEM/ mo l AVS )µ µ


9

on the extent to which a method was able to correctly classify a toxic sample as toxic and a non-

toxic sample as non-toxic. ROC curves were constructed by using acute toxicity data from 357

samples chosen from eight sources including freshwater and marine sediments. Species tested were

Hyalella azteca, Chironimus riparius, Neanthes arenaceodentata, Capitella capitata, Lumbriculus

variegates, Helisoma spp., Ampelisca abdita and Chironomus tentans. The results on the SEM/AVS

model evaluation showed that this approach has a very high sensitivity (96 %) i.e the extent to

which a model correctly classifies a toxic sample as toxic and is therefore protective of the

environment. Next to sensitivity both the negative and positive predictive capability was examined.

From this analysis it is clear that the SEM/AVS model provides an adequate negative predictive

power of 97 % but provides low positive predictive power of 55 %. Because the latter is the

likelihood that a sample exceeding the threshold is in fact toxic, it means that in a large number of

cases exceeding the SEM/AVS ratio does not result in any observed toxic effects. This is not

surprising because of the explicit exclusion of other binding phases in the approach, and since both

the SEM/AVS threshold of 1 and SEM-AVS threshold of zero are not intended to predict toxicity but

intended to define the circumstances where toxicity would not be expected.

As the evidence above is indicating the SEM/AVS concept does reasonably well in predicting the

absence of toxicity but is less usefull to predict metal related toxicity. In this regard using the

difference between the molar concentrations of SEM and AVS (SEM-AVS) instead of the SEM/AVS

ratio can already provide important information. The SEM-AVS difference gives insight into the extent

of additional binding capacity, the magnitude by which AVS binding has been exceeded, and when

organism response is considered, the potential magnitude of importance of other metal binding phases

(Hansen et al., 1996). At a molar SEM-AVS difference < 0 no effects are expected to occur.

2.3. DEALING WITH THE MULTIPLE METALS ISSUE

In applying the SEM-AVS model for a specific metal, such as nickel, it has to be taken into

consideration that ΣSEM represents the sum of different metals acting in a competitive manner when

binding to AVS. Acknowledging the existence of competitive displacement kinetics between various

metals and reduced sulfur, the SEM-AVS model can be made nickel specific. The procedure that is

used is to assign the AVS pool to the metals in the sequence of their sulfide solubility products.

Ranked from the lowest to the highest solubility product the following sequence is observed: SEMCu,

SEMPb, SEMCd, SEMZn and SEMNi. Meaning copper has the highest affinity for AVS, followed by


10

lead, cadmium, zinc etc until the AVS is exhausted. The remaining SEM is that amount present in

excess of the AVS. Mercury and silver have lower sulfide solubilities than nickel, but their molar

concentrations are always sufficiently low as to not influence the calculations.

To demonstrate, let � {SEMi} be the excess SEM for each of the ith metals. The least soluble

metal sulfide (of the five metals considered above) is copper sulfide. Thus, if the simultaneously

extracted copper is less than the AVS ({SEMCu} < {AVS}), then essentially all of it must be

present as copper sulfide with no additional SEMCu present, such that the difference between

SEMtotal and SEMCu = � {SEMCu} = 0. The remaining AVS binding pool is � {AVS} = {AVS}

- {SEMCu}. This computation is repeated next for lead and cadmium because these are the next

least soluble sulfides.

In the case of nickel, the excess SEMNi (µmol/gDW) can be calculated as follows:

Excess SEMNi = SEMNi - �AVSNi

�AVSNi = AVStot - (SEMCu + SEMPb+ SEMCd+ SEMZn) (if AVS exhausted = 0)

But even if there is an excess Ni this does not necessarily implies that toxicity will occur. The

SEM-AVS model can only be used to predict the absence of toxicity. The inability of the SEM-AVS

concept to predict toxicity is due, in part, to the neglect of the partitioning term.

SEM = AVS + KpCw

Where Cw represents a critical water toxicity value and Kp a partitioning coefficient.

Ditoro et al. (2005) extended the original concept in a kind of sediment Biotic Ligand Model in

which the BLM component adopts the EqP approach to relate toxicity of sediment-associated metals

to dissolved metal concentrations. The observation that metals such as nickel binds to organic carbon

suggests that organic carbon partitioning can be used as the most important partitioning phase to

predict the onset of toxicity ( (Mahony et al., 1996, Di Toro et al, 2005,US-EPA, 1999, US-

EPA, 2005). In this context it is assumed that toxicity occurs if the excess SEM goes beyond the

binding capacity of the organic carbon present in the sediment. Using this information it was shown

that the organic carbon normalized excess SEM can be used to predict toxicity):

SEMx, oc = ΣSEM − AVS

fOC


11

where fOC is the organic carbon fraction in the sediment.

For nickel the SEM,OC concentration is 1,100 µmol/g dry wt. Above this value toxicity is likely to

occur. Below no toxicity is expected.


12

3. OVERVIEW STUDIES INVESTIGATING THE EFFECT OF AVS ON NICKEL

TOXICITY

3.1. INTRODUCTION

Nickel was one of the first metals used to develop the SEM-AVS concept and literature is available

that support the use of the SEM/AVS concept for nickel. The tests performed on Ni included acute

and chronic laboratory and field exposures in both marine and freshwater sediments. Details of the

test design and main findings of the most relevant acute, chronic and field studies are described in

Table 1 and discussed hereunder.

Table 1: Overview Ni specific acute, chronic and field studies examining the influence of AVS on nickel

bioavailability

Reference Type of

study

Species Habitat Endpoint AVS model applicable

Ankley et al,

1991

Acute

(10d)

F

H. azteca epibenthic Survival YES

L. variegatus benthic Survival YES

Di Toro et al

1992

Acute

(10d)

M/F

A. abdita epibenthic Survival YES


Pesch et al,

1995

Acute

(10d)

M

N. arenaceodentata

benthic Survival YES

Doig and

Liber, 2006

Acute

(10d)

F


Acute

(10d)

F

H. azteca epibenthic Growth YES (in low AVS sediments/

high AVS sediments)

Vandegehuchte

et al, 2007

Chronic

(28d)

F

L. variegatus benthic Biomass YES if only SEM-AVS (0-1 cm)

Is considered.

No if SEM-AVS 0-3 cm is

used

Besser et al,

2011

Chronic (

at least

28d)

H. azteca epibenthic Survival,

growth,

biomass and

YES


13

F number of

offspring

G. pseudolimnaeus

epibenthic Survival,

growth and

biomass

YES

L. variegatus benthic Survival,

growth and

biomass

YES

T. tubifex benthic Survival,

growth,

biomass and

number of

offspring

Unbounded test results

C. dilutus benthic/epibenthic Survival,

growth,

biomass,

emergence

and egg

production

Unbounded test results

C. riparius benthic/epibenthic Unbounded test results

Hexagenia benthic/epibenthic Survival,

growth and

biomass

Toxic . effects occurs in situation

where SEM-AVS< 0. Could be due

to dietary contribution or formation

of oxic micro-habitats

L. siliquoidea Epibenthic Survival,

growth and

biomass

Not reliable test results

Boothman et

al, 2001

Field

M

Biological

indigenous

species

All habitats Abundance YES

Nguyen et al,

2011

Field

F

Biological

indigenous

species


Costello et al,

2011

Field F Biological

indigenous

species


3.2. ACUTE STUDIES

There are a number of short-term studies (10 days) available with regard to the applicability of the

SEM/AVS concept for nickel. These are outlined in the next sections.

3.2.1. ANKLEY ET AL., 1991


14

Ankley et al investigated the bioaccumulation and acute toxicity of cadmium and nickel in field-

collected contaminated sediments from Foundry Cove, which is at the edge of freshwater and

estuarine conditions in the Hudson river. At the time of sampling the sediments were freshwater.

Short term exposures (10 d) were conducted with the amphipod Hyalella azteca and the polychaete

Lumbriculus variegatus. Molar SEM/AVS rations ranged from less than one to greater then 200.

Samples with SEM/AVS ratios larger then 1 were consistently toxic for the amphipod whereas

sediments with ratios below 1 were not. L. variegatus was far less sensitive.

3.2.2. DI TORO ET AL., 1992

In the experiments of Di Toro et al (1992) short term (10 days) toxicity tests were conducted

with the marine amphipod Ampelisca abdita exposed to nickel spiked sediments. The final experiment

consisted of the exposure of the freshwater amphipod Hyalella azteca to field sediments contaminated

with cadmium and nickel from the Foundry Cove battery manufacturing facility. No mortality was

observed with SEM/AVS ratios below 1 demonstrating the predictive power of the ratio to classify

sediments as not acutely toxic due to cadmium and/or nickel contamination.

3.2.3. PESCH ET AL., 1995

In this study, the influence of AVS and interstitial water (IW) cadmium and nickel on bioavailability

and acute toxicity/bioaccumulation (10d) to the sediment-ingesting marine polychaete, Neanthes

arenaceodentata was investigated. Field sediments (Ninigret pond, 2 µmol AVS/g dry wt. and Long

Island Sound sediment, 18 µmol AVS/g dry wt ) were spiked with Ni or Cd to obtain nominal

SEM/AVS ratios of 0, 0.1, 0.3, 1, 3, 10, 30, 100. For the spiked sediments, measured SEMNi

concentrations in the sediments ranged from < 0.3 to 550 µmol/g dry wt. In addition sediment

toxicity tests were also conducted with 16 contaminated field sediments from Foundry Cove. In these

experiments SEM/AVS ratios ranged between < 0.04 to 125 with SEMNi concentrations (< 0.28-

449 µmol/g dry wt). During the experiments AVS, SEM, IW metal concentrations, mortality,

burrowing and tissue concentrations of the worms were determined. Water-only experiments were also

conducted


15

All data were normalized to AVS and IW toxic units (IWTU = IW metal concentration/LC50 value

of water-only exposure). In all experiments, no significant mortality was observed when SEM/AVS

ratios were < 1 and IWTU were < 1. The nickel-spiked experiments, when SEM/AVS ratios or

IWTUs were > 1, sediments were either lethal to the worms or the worms did not burrow. The

absence of toxicity to the polychaetes was correctly predicted with the SEM/AVS concept (no toxicity

when SEM/AVS < 1).

3.2.4. DOIG AND LIBER, 2006

In this study 10-d nickel spiked sediment tests using the amphipod H. azteca were conducted using

four freshwater sediments varying substantially in AVS and OC content. AVS ranged from 0.73 to 44

µmol/g dry wt. Generally, lethal and non-lethal toxicity test endpoints were reasonable predictable in

low AVS sediments. Predictions on the combined protective effect of AVS and OC were

overestimated in sediments containing mid to high AVS concentrations (27.9-44 µmol.g dry wt.). It

was hypothesized by the authors that a lack of equilibrium between the spiked nickel and the

associated NiS could have likely resulted in the presence of soluble Ni species such as nickel

bisulphide complexes potentially contributing to the observed toxic effects.

3.3. CHRONIC STUDIES

3.3.1. VANDEGEHUCHTE ET AL., 2007

In the study of Vandegehuchte et al (2007) it was examined for the first time if the SEM-AVS

concept that was successfully applied to acute studies could be extended to predict the absence of

chronic Ni toxicity to the oligochaete deposit-feeding worm Lumbriculus variegatus and (2) if the

organic carbon normalized excess SEM; i.e. [SEM-AVS]/fOC predicts the magnitude of Ni toxicity to

L. variegatus. A 28-day toxicity experiment was performed in which biomass production of L.

variegatus was determined in two natural sediments with different [AVS] and fOC, spiked at different

Ni concentrations. The absence of toxicity was predicted correctly by the [SEM-AVS] < 0 criterion

when only the 0-1 cm surface layer of the sediment was considered, but not when the whole bulk

sediment was taken into account (0-3 cm). In both sediments, the same [SEM-AVS]/fOC at the


16

surface corresponds with a similar decrease in L. variegatus biomass. Thus, [SEM-AVS]/fOC in the

surface layer accurately predicted the magnitude of toxicity. This measure was therefore deemed a

good estimator of toxicologically available Ni. On the other hand, the free Ni2+ ion activity in the

overlying water appeared to be an equally good predictor of the magnitude of toxicity. Consequently,

it was not possible to determine the relative importance of the overlying water and pore water

exposure route with the semi-static laboratory experiments leading to an unintended artifact, i.e.

release of nickel from the spiked sediment to the overlying water compartment blurring the final

conclusions. The findings of the Vandegehuchte et al (2007) triggered additional research (Besser

et al., 2011) with adapted spiking procedures avoiding a release of nickel to the overlying water.

3.3.2. BESSER ET AL, 2011.

Besser et al (2011) evaluated the relative sensitivity of invertebrate taxa to toxic effects of two

nickel-spiked sediments: sediment from the Spring River, Missouri, which had low concentrations of

the important metal-binding components, total organic carbon (TOC) and AVS; and sediment from

West Bearskin Lake, Minnesota, which had high TOC and high AVS. Eight taxa were tested in

flow-through sediment exposure systems with automated replacement of overlying water: two

amphipods, Hyalella azteca and Gammarus pseudolimnaeus; two midges, Chironomus dilutus and

Chironomus riparius; two oligochaetes, Lumbriculus variegatus and Tubifex tubifex; a mayfly,

Hexagenia sp.; and a freshwater mussel, Lampsilis siliquoidea. These tests lasted at least 28 days

and included multiple chronic toxicity endpoints (survival, growth, and biomass for all eight taxa;

adult emergence and egg production for Chironomus spp.; and number of offspring for Hyalella

azteca and Tubifex tubifex) to determine the most sensitive responses of each species. The three

most sensitive taxa (plus Tubifex) were tested with six additional sediments that represented a

gradient of physicochemical characteristics, including AVS, TOC, and particle size distribution. Toxicity

of nickel-spiked sediments to the amphipods, Hyalella and Gammarus, was consistent with the

hypothesis that AVS is a primary control on pore-water nickel concentrations and on toxicity of nickel

in sediments. For these taxa, nickel-spiked sediments were not toxic if nickel concentrations were

less than AVS concentrations on a molar basis. In contrast, toxic effects on the burrowing mayfly

Hexagenia occurred in several sediments with nickel concentrations less than the theoretical AVS

binding capacity. These divergent results could indicate that AVS does not strongly control nickel


17

bioavailability to Hexagenia, perhaps as suggested by the authors because ingestion of sediment

particles was an important route of nickel exposure for this species. Alternatively, it was possible that

the sampling methods used in this study did not adequately measure localized concentrations of

AVS or pore-water nickel (or both) in the burrows inhabited by Hexagenia. The observations

resulted in the development of bioavailability models based on AVS for the amphipods H. azteca and

G. pseudolimnaeus and the mayfly Hexagenia.

3.4. FIELD STUDIES

For most metals the applicability of the SEM-AVS model has been proven primarily in laboratory

settings. For nickel three studies investigated the utility of the SEM-AVS model under field

conditions. (Boothman et al, 2001, Nguyen et al, 2011 and Costello et al, 2011). These studies

cover mainly streams and a range of different sediment types, with varying AVS concentrations, were

conducted during different seasons, and were carried out in different geographical locations (Europe

and North America) and in different types of systems (lotic and lentic), with varying water quality

and abiotic parameters. In abovementioned colonisation studies, the abundances of major taxa

(including classes and families) which colonised the initially defaunted sediments, were studied. The

field work also included determinations of abundances down to genera and species. All colonisation

studies included a variety of organisms differing in morphology and (feeding and burrowing)

behaviour and thus expected to differ in exposure and sensitivity.

3.4.1. BOOTHMAN ET AL., 2001

A 4 month large scale field recolonisation study was conducted by Boothman et al. (2001) in

Narragansett Bay, a temperate marine system. Sediments (77 % silt/23 % sand, 1.2 % organic

carbon and initial AVS concentration of 9 µmol/g dry wt.) were spiked with equimolar quantities of

cadmium, copper, lead, nickel and zinc. Nominal total metal/AVS ratios were 0.1, 0.8 and 3.0.

On day 119 of the experiment, sediment cores from the different treatments were collected and

sectioned to allow profiles of metals and AVS to be determined with finer resolution. Overall,

concentrations of metals in the IW were undetectable when SEM/AVS < 1 and were enhanced when


18

SEM/AVS > 1, with the least soluble sulfides always below detection limits. Based on SEM/AVS

data, it was predicted that metals could be present in interstitial waters of the 0.8x and 3.0x

treatments. Metals were measured in the interstitial water of the 3.0x treatment and decreased with

time. However, after 119 days exposure faunal assemblages were similar to those in the control.

Lack of biological response was related to vertical distributions of AVS and SEM, with AVS

exceeding SEM even in the highest concentration in the surficial sediment sections. Because benthic

species initially colonize sediments at the seawater interface, conditions at the interface would control

colonization and effects on benthic organisms more than the deeper sediment conditions. Since metals

are released from the uppermost sediment layer and readily washed away, this may have explained

the lack of biological effects.

The results from Boothman et al (2001) suggest that when metals-spiked sediments are placed in

natural dynamic environments, field conditions do not allow high concentrations of excess metals to

exist in the surficial sediments for long periods, most likely due to physical resuspension releasing

metals from pore water into overlying waters This emphasizes again the merit of using fine-scale

vertical gradients and seasonal variations when using AVS and SEM measurements to assess the

bioavailability of metals under natural conditions.

3.4.2. NGUYEN ET AL, 2011

Nguyen et al (2011) performed field experiments in four freshwater systems to assess the effects

of nickel on the benthic macro-invertebrate communities. Sediments were collected from four

uncontaminated study sites (in Belgium, Germany, and Italy), spiked with nickel, and returned to

the respective field sites. The colonization process of the benthic communities was monitored during

a nine-month period. Nickel effect on the benthos was also assessed in the context of equilibrium

partitioning model based on acid volatile sulfides (AVS) and simultaneously extracted metals

(SEM). Benthic communities were not affected at (SEM - AVS) < 0.4 mmol/g, (SEM -

�AVS)/fraction of organic carbon (ƒOC) < 21 mmol/g organic carbon (OC). Sediments with

(SEM -� AVS) > 2 mmol/g, (SEM - AVS)/ƒOC > 700 mmol/g OC resulted in significant

adverse effects on the benthic macro-invertebrate community.

3.4.3. COSTELLO ET AL, 2011


19

Five lotic sediments with a range of sulfide and organic carbon contents were spiked with four

concentrations of nickel, deployed in streams for eight weeks, and examined for colonizing macro-

invertebrates. An important difference with previous field studies is that a new spiking technique has

been used in which all sediments were amended with NiCl2 by a “superspike” method, equilibrated

for four weeks and then diluted with untreated sediment to achieve the desired concentrations

avoiding rapid loss of nickel to the overlying water. Previous field studies included a fraction of Ni

that was loosely bound and rapidly lost when placed in the field. This loosely bound fraction needs

to be accounted for in the interpretation of the results of these earlier studies.

After four weeks, colonizing macro-invertebrates showed a strong negative response to the Ni-treated

sediments and SEM-AVS models of bioavailability differentiated between toxic and nontoxic conditions.

By Week eight, relationships deteriorated between colonizing macro-invertebrates and SEM-AVS model

predictions. Total nickel in the sediment did not change through time; however, nickel partitioning

shifted from being dominated by organic carbon at deployment to associations with iron and

manganese. Combined geochemical and toxicity results suggest that iron and manganese oxides in

surface sediments resulted in nickel being less available to biota. This implies that current SEM-AVS

models may overestimate bioavailable nickel in sediments with oxic surface layers and sufficient iron

and manganese.

3.5. CONCLUSION ENDPOINT TOXICITY

The review of the existing nickel specific literature shows that the SEM/AVS concept has proven to

be adequate in predicting the absence of toxicity (for which it was originally developed) of nickel in

both short-term and long-term experiments, with spiked sediments as well as field sediments.

Indeed almost all of the sediments tested did not show adverse effects on benthic biota as long as

SEM/AVS levels were below one. Situations where toxicity was observed while according to the

SEM-AVS model no metals should be bioavailable (SEM/AVS <1) does not automatically imply that

the SEM/AVS model is flawed. In those cases it was demonstrated that this observation could be

linked to the lack of sampling resolution when using only the mean measured SEM-AVS values

since these values do not always reflect what a benthic organisms actually 'see’ in the surficial

sediment layers. In any case if toxicity did occur in field sediments care should also be taken in the


20

interpretation if this toxicity could really be contributed to metals or that the toxic effects are due to

the presence of other contaminants that have not been measured.

In case the SEM/AVS ratio is above one, it is possible that metals (and in this case nickel) can

be present in sediment pore water. The extent to which metals in pore water will be toxic is

dependent upon 1) the metal concentration and 2) the presence of other ligands in the pore water

that may mitigate nickel toxicity. This does not necessarily imply that nickel toxicity should be

observed. The latter is dependent on the magnitude of exceedance meaning that toxicity will only

occur when critical effect levels (LOEC values, LC50 values) are reached.

Factors controlling metal toxicity at SEM-AVS levels > 1, are less well defined. In oxygenated

sediments these factors include sorption to iron and/or manganese oxides, clay minerals and

sediment organic matter (ITRC, 2011).

However, even under reduced conditions metal sulfide phases have been demonstrated to be

bioavailable to infaunal organisms. The sometimes-poor correlation observed between benthic

invertebrate tissue data and SEM-AVS predictions of metal bioavailability likely results from direct

metal assimilation through ingestion (Lee et al, 2000). The main subject of the next chapter is to

elucidate the mechanisms of uptake and bioaccumulation of metals and nickel in particular under

conditions that SEM-AVS <0 and its consequences to use the SEM-AVS model in a risk

assessment context.


21

4. OVERVIEW STUDIES INVESTIGATING THE EFFECT OF AVS ON NICKEL

BIOACCUMULATION

4.1. INTRODUCTION

The SEM-AVS model has been validated for both acute and chronic toxicity and the SEM-AVS

model predicts that no toxicity should occur when excess AVS is present.. Studies examining the

bioaccumulation of metals in anaerobic sediments showed in general that in most of the cases metal

accumulation is reduced when SEM-AVS < 1 (Ankley et al, 1996b). However, in some case

bioaccumulation was best correlated with total metal content in the sediment irrespective of the AVS

content (Lee at al, 2000, De Jonge et al, 2009, De Jonge et al, 2010).

The most relevant studies describing this specific phenomenon for nickel are discussed here below.

4.2. BIOACCUMULATION STUDIES

4.2.1. ANKLEY ET AL., 1991

Ankley et al investigated the bioaccumulation of cadmium and nickel in contaminated estuarine

sediments. The amphipod Hyalella azteca and the polychaete Lumbriculus variegatus were exposed to

cadmium and nickel spiked sediments in short term exposures (10d). Bioaccumulation in the

polychaetes was significantly elevated when SEM/AVS were > 1 but small metal concentrations were

also observed also at SEM/AVS ratios < 1.

4.2.2. PESH ET AL, 1995

A description of the test design can be found in section 4.2.3. Next to the toxicity endpoints Pesch

et al (1995) described the bioaccumulation of nickel in the marine worm Neanthes arenaceodentata.

In all three experiments, accumulation of metals by worms in sediments with SEM/AVS ratios greater

than 1.0 was generally higher than those with sediments with ratios less than 1.0 but metals were

accumulated by worms in cadmium and nickel-spiked sediments with SEM/AVS ratios < 1. This


22

phenomenon was only observed in the laboratory spiked sediments and did not occur in the field

sediments. The presence of nickel in worms from sediments with SEM/AVS < 1, was suggested to

be resulting from release of nickel from oxidized metal sulfide (as a result of burrowing), uptake of

metal from ingested sediment, or adsorption to body surfaces.

4.2.3. DE JONGE ET AL, 2009

De Jonge et al (2009) evaluated the influence of AVS on the accumulation of

sediment-bound metals in two widespread benthic invertebrates (Chironomus and Tubifex

species) collected from the field. In total 17 historicall polluted Flemish lowland rivers

were selected on containing high metal concentrations in the sediment but low

concentrations in the overlying water, to minimize possible accumulation of the

invertebrates from the surface water. Metal accumulation in tubificids generally followed the

same accumulation pattern as in the chironomids. At several sampling sites a high

amount of metals (in particlaur Pb) was accumulated even when SEM-AVS < 0. In

general metal accumulation was most strongly correlated with total metal concentrations in

the sediment and sediment metal concentrations normalized for LOI and clay content.

4.2.4. DE JONGE ET AL, 2010.

In a follow up study De Jong et al (2010) examined if the relationship between

steady state tissue concentrations and relevant sediment phases under SEM-AVS <

1conditions still hold when comparing invertebrates with different feeding behaviour and

ecological preferences. Natural sediments together with benthic and epibenthic invertebrates

were sampled at 28 Flemish lowland rivers. The study confirmed that tissue

concentrations in truly benthic taxa was primarily influenced by total metal concentrations

in the sediment normalized for LOI and dissolved metal concentration in the surface

water. In addition for most metals a good correlations was found between SEMMe and

observed steady state tissue concentrations. For some metals (Zn, Cr, Cu and Cd)

SEMMe –AVS was also positively correlated with the tissue concentrations found in some


23

benthic species (Asellus. aquaticus, Erpobdella octoculata). However, concerning Co, Ni,

As and Pb no significant correlation between tissue concentrations and SEM-AVS were

found. The benthic taxa accumulated high amounts of Cr, Co, Ni, Zn, As and Cd even

when AVS largely exceeded SEM. The epibenthic taxa did not accumulate Cu, Cr, Cd,

Zn and Pb when SEM-AVS <0. However, regarding Co, Ni, As and Ag some

accumulation was noticed at negative SEM-AVS values. However, all invertebrates

including the epibenthic species accumulated Ni even if SEM-AVS < 0.

4.2.5. DE JONGE ET AL, 2011

De Jonge et al (2011) tried to link in more detail metal binding sediment characteristics like AVS,

metal bioaccumulation with the internal metal distribution in the benthic oligochaete Tubifex tubifex in

order to clarify if accumulated sulfide bound metals can be actually harmful to the affected organism.

For that purpose 15 streams were sampled and both the total as the subcellular compartmentalization

of metals within the field collected tubificids were determined.

High metal concentrations were measured in the whole tissue of the collected tubificids compared to

the levels normally found in clean conditions. The results showed that Cd, Pb, Ni and Cr were

mainly stored as biological detoxified metals (BDM) such as metal rich granules and heat stable

proteins, while Cu, Zn, As and Ag were mostly available in the metal sensitive faction (MSF) such

as organelles (nuclear, mitochondrial and microsomal fractions) and heat denaturated proteins. .

When SEM exceeded AVS, detoxification mechanisms became saturated and accumulation in the MSF

started to occur, resulting in an enhanced MTLP (metallothionein like proteins) production.

4.3. DISCUSSION AND CONCLUSION

It should be noted that for sediment risk assessment purposes, measuring relationships between tissue

concentrations and SEM/AVS doesn’t really mean that much.. The mechanisms of Ni toxicity for

invertebrates is mainly an ionoregulatory mechanism. (Pane et al, 2003, Keithly et al, 2004).

Toxicity does not depend on total accumulated metal concentration but is related to a threshold

concentration of internal metabolically available metal (Rainbow, 2007). Toxicity ensues when the


24

rate of metal uptake from all sources exceeds the combined rates of detoxification and excretion of

the metal concerned. Many of the studies mentioned above explored the relationships between tissue

metal concentration and sediment concentration. Only one examined detoxification mechanisms in more

detail.

The observation that in some studies bioaccumulation in benthic invertebrates, can occur even when

SEM-AVS < 0 can be partly ascribed to the fact that in some studies regarding bioaccumulation,

the organisms were not purged before metal analysis (Ankley et al, 1996). The presence of

sediment bound metals not absorbed in the gut could lead to an overestimation of the actual metal

accumulation. This is for example the case in the Pesch et al (1995) study where the polychaete,

Neanthes arenaceodentata accumulated nickel at SEM/AVS ratios < 1. This phenomenon was only

observed in the laboratory spiked sediments and did strangely enough not occur in the contaminated

field sediments. The nickel and cadmium concentration used in the Pesch et al (1995) study was,

at the higher end, (200 and 700 mg/g dry wt., respectively at SEM-AVS < 1) making it possible

that even minimal contributions from residual gut contents and/or surface absorption could contribute

significantly to the total body burden of nickel or cadmium measured in the polychaete (Ankley et

al., 1996).

However, even if data biased by this confounding factor are omitted from the database there is a

enough evidence (mainly for marine invertebrates) indicating that feeding behavior and dietary uptake

next to porewater may play an important role in metal bioaccumulation (Wang et al, 1999, Lee et

al, 2000, Lee et al, 2001).

Results of Doig and Liber (2006) indicated that, even when SEMNi –AVS < 0 bioavailable Ni

concentrations can still occur in the sediment’s pore water due to the lack of equilibrium between

precipitated sulphides and dissolved metal species. As a result reduced growth for the amphipod H.

azteca was observed in their experiment even in the presence of excess AVS. However, enhanced

bioaccumulation can be best explained by the fact that benthic invertebrates ingest sediment particles

as their main food source, regardless of AVS (Lee et al 2000, 2001). Bioavailability of

sedimentary contaminants will depend upon solubilisation under digestive biochemical conditions. As

conditions within animal guts can differ substantially from sedimentary environments, metals availability

can be modified drastically. For example digestive processes and gut conditions such as Eh, pH and

enzyme or surfactant activity can affect the release of ingested metals in the gut and control uptake

(Griscom et al, 2002, Mayer et al, 1996/1997, Chen and Mayer, 1999).


25

The more sediment ingestion becomes the dominant metal exposure route, the more metal

accumulation in the organism seemed not to be controlled by AVS concentrations in the sediment as

demonstrated by De Jonge (2010) who observed that in general AVS concentrations were generally

better correlated with metal accumulation in the epibenthic invertebrates indicating that the relation

between AVS and metal accumulation in aquatic invertebrates is highly dependent on feeding

behaviour and ecology.

The observation that metals can be taken up under SEM-AVS conditions < 0 is not in contradiction

with the overall SEM-AVS model or with our general understanding of metal bioaccumulation. Figure

2 gives a general overview of the processes involved.

Figure 2: Conceptual diagram for evaluating bioavailability processes and bioaccessibility for metals in

soil, sediment or aquatic systems (adapted scheme from US-EPA, 2007)

The public notion that metal bioavailability in sediments is controlled only by geochemical equilibration

of metals between pore water and reactive sulfides” is incorrect. The absence of metals in pore

water does not mean that organisms should not be able to accumulate metals from other sediment

phases or other exposure routes (e.g. dietary route). It merely indicated that the toxicologically

relevant species, Me2+, should not be available in sediment pore water. These are two very

different things and for risk assessment purposes should be clearly separated. The relevance of


26

dietborne nickel toxicity has recently been published (Evens at al, 2012) and specific research

programs has been established to explore this subject in more detail (Nguyen et al, 2012, Burton

et al, 2012).

The main question to be solved from a risk assessment point of view is to which extent

accumulated sulphide-bound metals can be actually harmful to the affected organism. This because

the amount of accumulated metal concentration is not necessarily related to the development of toxic

effects. Rainbow (2002 and 2007) already suggested that the onset of toxic effects depends only

on the concentration of accumulated metals that are present in a metabolically available form.

A better understanding of the internal compartmentalization of metals in organisms is key to

understand its consequences for toxicity. In their review Vijver et al (2004) clearly demonstrated

that aquatic organisms are able to control metal concentrations in certain tissues of their body to

minimize damage of reactive forms of essential and non-essential metals. The metabolic pathways

and internal compartmentalization of metals is metal specific and can be ascribed to their specific ion

radius and electronegativity. Metals such as nickel and zinc have no binding preference and will form

ligands with many functional groups. Dumas and Hare (2008) showed for two common types of

benthic invertebrates (Chironomus riparius and Tubifex tubifex) that at least half of the accumulated

nickel was present in fractions that are purportedly detoxified (granules and metal binding proteins).

No information was available on the AVS content of the test sediments.

The biological significance of the accumulated metal concentrations under SEM-AVS conditions < 0

will depend on the way organisms cope with the increased metal exposure. As demonstrated with the

results of De Jonge et al (2011) for nickel there was a remarkable difference in the subcellular

distribution of accumulated Cd, Ni and Co between anoxic and oxic sediments. Nickel distribution

changed from a major storage as BDM under anoxic conditions (SEMNi-AVS< 0) to an equal

distribution over BDM and MSF when SEM exceeded AVS. These results indicate that metal uptake

through the ingestion of sulphide-bound particles is mainly sequestered as BDM and hence detoxified

and little available to the metabolism of T. tubifex under anoxic conditions. Biomagnification of nickel

has not been further investigated in the current report. A detailed discussion on the relevance of

biomagnification for nickel can be found in section 3.2.4.2 of the nickel EU RAR report (EU,

2008). The evidence presented in that section suggests that Ni does not biomagnify, but rather that

it tends to exhibit biodilution, particularly when upper levels of the food chain are considered.


27


28

5. OVERALL CONCLUSION

The recent body of work shows how the binding strength of reactive acid volatile sulfides (AVS)

controls pore water concentrations of divalent metals such as nickel. In these studies, AVS

concentrations consistently limit metal toxicity to invertebrates in sediments where molar AVS

concentrations exceed molar concentrations of extractable metals.

Effects of AVS on metal bioaccumulation are somewhat more ambiguous. Although, the preponderance

of studies indicate indeed reduced accumulation of metals at sediment metal/AVS ratios of less than

1 (Ankley et al, 1996b), there were exceptions to this general observation (Pesch et al, 1995,

Lee et al 2000a, 2000b, De Jonge et al 2009, 2010 and 2011) questioning the notion that metal

bioavailability in sediments is controlled only by geochemical equilibration of metals between pore

water and reactive sulfides.

The dietary route seems to play an important role in explaining the observations that metals are

taken up from the sediment irrespective of the AVS concentration. It should, however, kept in mind

that bioaccumulation does not represent a toxicological effect and an unambiguous connection between

observed levels of accumulation and effects has not yet been made. At SEM-AVS concentrations <

0 metals extracted in the gut from the ingested metal sulfides are detoxified and stored in granules

while at an overload of the AVS system metals can be found in a more easily accessible pool.

Overall the recent results support the tenet that AVS controls metal bioavailability in particular with

relation to chronic effects and can therefore be used in a risk assessment framework.


29

6. REFERENCES

Ankley G.T., Phipps G.L., Leonard E.N., Benoit D.A., Mattson V.R., Kosian P.A., Cotter A.M.,

Dierkes J.R., Hansen D.J. and Mahony D.J., 1991. Acid Volatile sulfide as a factor mediating

cadmium and nickel bioavailability in contaminated sediments. Environmental Toxicology and Chemistry,

10, 1299-1307.

Ankley G.T., D.M. Di Toro, D.J. Hansen and W.J. Berry, 1996a. Technical basis and proposal for

deriving sediment quality criteria for metals. Environmental Toxicology and Chemistry, 15 (2): 2056-

2066.

Ankley, G.T., 1996. Evaluation of metal/acid-volatile sulfide relationships in the prediction of metal

bioaccumulation by benthic macroinvertebrates. Environmental Toxicology and Chemistry, 15 (12),

2138-2146.

Besser, J.M., W.G. Brumbaugh, N.E. Kemble, C.D. Ivey, J.L. Kunz, C.G. Ingersoll and D. Rudel,

2011. Toxicity of nickel-spiked freshwater sediments to benthic invertebrates: spiking methodology,

species sensitivity and nickel bioavailability. USGS Open-file Report Series.

Birch, GF, Taylor S.E. and Matthai C., 2001. Small-scale spatial and temporal variance in the

concentration of heavy metals in aquatic sediments: a review and some new concepts. Environmental

pollution 113 (3) 357-372.

Buykx S.E.J. & van den Hoop M.A.G.T. 1996. The simultaneous influence of pH and Acid Volatile

Sulphide on the speciation of heavy metals in calcareous sediments. Report no. 502501043. National

institute of public health and the environment, Bilthoven, The Netherlands.

Burton G.E., Jr, Nguyen T.H., Janssen C., Baudo R ., McWilliam R., Bossuyt B., Beltrami M. and

Green A., 2005. Field validation of sediment zinc toxicity. Envirinmental Toxicology and Chemistry,

24 (3) 541-553.

Burton G., A. Green, R. Baudo, V. Forbes, L. Nguyen, C. Janssen, J. Kukkonen, M. Leppanen,

L. Maltby, A. Soares, K. Kapo, P. Smith and J. Dunning., 2007. Characterizing sediment acid

volatile sulfide concentrations in European streams. Environmental Toxicology and Chemistry, 26 (1),

1-12.

Burton et al, 2012. Research program in progress.


30

Costello, D.M., Burton G.A., Hammerschmidt C.R., Rogevich E.C. and Schlekat C.E, 2011. Nickel

phase partitioning in field-deployed sediments. Environ Sci. Technology, 45 (13), 5798-57805

De Jonge M., Blust R. and Bervoets L., 2010. The relation between Acid Volatile Sulfides (AVS)

and metal accumumation in aquatic invertebrates : implications of feeding behavior and ecology.

Environmental Pollution 158, 1381-1391.

De Jonghe M., M. Eyckmans, R. Blust and L. Bervoets, 2011. Are accumulated sulfide-bound

metals metobolically available in the benthic ologochaete Tubifex tubifex. Environ. Sci. Technol., 45,

3131-3137.

DeWit, T.H., R.C. Swart, D.J. Hansen, W.J. Berry and D. McGovern, 1996. Interstitial metal and

acid volatile sulfide predict the bioavailability of cadmium during a full life-cycle sediment toxicity test

using the estuarine amphipod, Leptocheirus plumulosus. Environmental Toxicology and Chemistry, 15-

2095-2101.

Di Toro, D.M., J.D. Mahony, D.J. Hansen , K.J. Scott, M.B. Hicks, S.M. Mays and M.S.

Redmond,1990. Toxicity of cadmium in sediments: the role of acid volatile sulfides. Environmental

Toxicology and Chemistry, 9, 1487-1502.

Di Toro D.M., J.D. Mahony, D.J. Hansen, K.J. Scott, A.R. Carlson and G.T. Ankley 1992. Acid

volatile sulfide predicts the acute toxicity of cadmium and nickel in sediments. Environmental Science

and Technology, 26, 96-101.

Di Toro D.M., J.A. McGrath, D.J. Hansen, W.J. Berry, P.R. Paquin, R.Mathew, K.B. Wu, R.C.

Santore, 2005. Predicting sediment metal toxicity using a sediment Biotic Ligand Model: methodology

and initial application. Environmental Toxicology and Chemistry, 24 (10), 2410-2427.

Doig L.E. and Liber K., 2006. Influence of dissolved organic matter on nickel bioavailability and

toxicity to Hyallela azteca in water-only exposures.

Dumas J. and Hare L., 2008. The internal distribution of nickel and thallium in two freshwater

invertebrates and its relevance to trophic transfer. Environ. Sci. Technol., 42 (14).

EU RAR, 2008. European Union Risk Assessment Report Nickel and Nickel compounds.

Evens R., De Schamphelaere K., Balcaen L., Wang Y., De Roy K., Resano M., Florez M., Boon

N., Vanhaecke F. and Janssen C., 2012. The use of liposomes to differentiate between the effects


31

of nickel accumulation and altered food quality in Daphnia magna exposed to dietary nickel. Aquatic

Toxicology 109, 80-89.

Grabowski L.A., J.L.J. Houpis, W.I Woods,. and K.A. Johnson, 2001. Seasonal bioavailability of

sediment-associated heavy metals along the Mississippi river flood plain. Chemosphere 45, 643-651.

Griscom SB, Fisher NS and Luoma SN. 2000. Geochemical influences on assimilation of sediment-

bound metals in clams and mussels. Environ. Sci. Technol., 91-99.

Griscom SB., Fisher NS, Aller RC and Lee BG. 2002. Effects of gut chemistry in marine bivalves

on the assimilation of metals from ingested sediment particles.

Hare L., Carignan R. and Huerta-Diaz M. 1994, A field study of metal toxicity and accumulation by

benthic invertebrates-implications for the acid-volatile sulphide (AVS) model. Limnol. Oceanogr., 39,

1653-1668.

Howard D.E., Evans R.D., 1993. Acid-volatile sulfide (AVS) in a seasonally anoxic mesotrophic

lake: Seasonal and spatial changes in sediment AVS. Environmental Toxicology and Chemistry

12:1051-1057.

ITRC, 2011. Incorporating bioavailability consideration into the evaluation of contaminated sediment

sites. Technical/regulatory guidance. Interstate Technology and Regulatory Council Contaminated

Sediments Team. February, 2011.

Keithly, J, Brooker JA, DeForest DK, Brix KV, Wu BK (2004): Acute and chronic toxicity of nickel

to cladoceran (Ceriodaphnia dubia) and amphipod (Hyalella azteca). Environmental Toxicology and

Chemistry, 23 (3), 691-696.

Lee B.G , S.B Griscom, J.S. Lee, H.J. Choi, C.H. Koh, S.N. Luoma and N.S. Fischer, 2000.

Influences of dietary uptake and reactive sulfides on metal bioavailability from aquatic sediments.

Science, 287, 282-284.

Liber K., D. Call, T. Markee, K. Schmude, M. Balcher, F. Whiteman and G.T. Ankley, 1996.

Effects of acid volatile sulfide on zinc bioavailability and toxicity to benthic macro invertebrates in

spiked-sediment field experiments. Environmental Toxicology and Chemistry,15, 2113-2125.

Mayer L.M., Chen Z., Findlay R.H., Fang J., Sampson S., Self R.F., Jumars P.A., Quetel C. and

Donard O.F.., 1996. Bioavailability of sedimentary contaminants subject to deposit feeder digestion.

Environ.Sci. Technol., 30, 2641-2645.


32

Mayer L.M., Schick L.L., Self R.F.L., Jumars P.A., Findlay R.H., Chen.Z. and Sampson S., 1997.

Digestive environments of benthic macroinvertebrate guts: enzymes, surfactants and dissolved organic

matter. Journal of Marine Research, 55, 7854-812.

Nguyen L.T., G.A. Burton Jr, C.E. Schlekat and C.R. Janssen, 2011. Field measurement of nickel

sediment toxicity: role of acid volatile sulfide. Environmental Toxicology and Chemistry, 30 (1), 162-

172.

Ngyuen et al, 2012. Research program in progress.

Pane, E.F., Smith, C., McGeer, J.C., Wood, C.M., 2003b. Mechanisms of acute and chronic

waterborne nickel toxicity in the freshwater cladoceran, Daphnia magna. Environmental Science and

Technology 37, 4382-4389.

Pesh, C.E., D.J. Hansen, W.S. Boothman, W.J. Berry and J.D. Mahony, 1995. The role of acid

volatile sulfide and interstitial water metal concentrations in determining bioavailability of cadmium and

nickel from contaminated sediments to the marine polychaete Neanthes arenaceodentrata.

Environmental Toxicology and Chemistry, 14: 129-141.

Peterson G.S., Anklet G.T. and Leonard E.N, 1996. . Effect of bioturbation on metal-sulfide

oxidation in surficial freshwater sediments. Environmental Toxicology and Chemistry, 15 (12), 2147-

2155,

Poot A, Gillssen F, Kolemans AA, 2007. Effects of flow regime and flooding on heavy metal

availability in sediment and soil of a dynamic river system. Environmental Pollution 148/3:779-787.

Prica M., Dalmacija B., Roncevic S., Krcmar D. and Becelic M., (2007). A comparison of

sediment quality with acid volatile sulfide (AVS) and simultaneously extracted metals (SEM) ratio in

Vojvodina (Serbia) sediments. Science of the Total Environment, 389, 235-244.

Rainbow S., 2007. Review article: Trace metal bioaccumulation: models, metabolic availability and

toxicity. Environment International 33, 576-582.

Simpson S.L. 2001. A rapid screening method for acid-volatile sulfide in sediments. Environmental

Toxicology & Chemistry. 20 (12), 2657-2661.

US-EPA, 1999. Integrated approach to assessing the bioavailability and toxicity of metals in surface

waters and sediments, EPa number 822E99001. 348 pages.


33

US-EPA, 2005. Procedures for the derivation of Equilibrium Partitioning Sediment Benchmarks

(ESBs) for the protection of benthic organisms: metal mixtures (cadmium, copper, lead, nickel,

silver and zinc). EPA/600/R-02/011.

US-EPA, 2007. Framework for Metals Risk Assessment. EPA 120/R-07/001 March 2007.

Van Den Berg, G.A., Loch J.P., Van Der Heijdt L.M. and Zwolsman J.G., 1998. Vertical

distribution of acid volatile sulfide and simultaneously extracted metals in a recent sedimentation area

of the river Meuse in the Netherlands. Environmental Toxicology and Chemistry, 17, 4, 758-763.

Van Den Berg, G.A., Loch J.P., Van Der Heijdt L.M. and Zwolsman J.G., 1998. Vertical

distribution of acid volatile sulfide and simultaneously extracted metals in a recent sedimentation area

of the river Meuse in the Netherlands. Environmental Toxicology and Chemistry, 17, 4, 758-763.

Van Den Berg, G.A., Meijers G.G., Van Der Heijdt L.M. and Zwolsman J.J., 2001. Dredging

related mobilisation of trace metals: a case study in the Netherlands. Wat. Res., 35 (8), 1979-

1986.

Van den Berg, G.A, Buykx, S.E.J., Van Den Hoop, M.A.G.T., Van Der Heijdt L.M. and Zwolsman

J.J.G., 2001. Vertical profiles of trace metals and acid-volatile-sulphide in a dynamic sedimentary

environment: Lake Ketel, The Netherlands. Applied Geochemistry, 16, 781-791.

Van Den Hoop, M, Den Hollander H. and Kerdijk H.N., 1997. Spatial and seasonal variations of

acid volatile sulphide (AVS) and simultaneously extracted metals (SEM) in Dutch marine and

freshwater sediments. Chemosphere, 35, 10, 2307-2316.

Vangheluwe M, (2005) Probabilistic assessment of copper bioavailability in sediments. Final report.

Vangheluwe M., Verdonck F., Heijerick D., Shtiza A., Burtin A., Taulbee K., Boucard T., Mäkinen

J. and Gonzales E., 2008. Acid volatille sulphide survey of large river and/or lake sediments in the

United Kingdom (England and Wales) Spain and Finland in 2007. Final report september 2008.

Van Griethuysen C, Luitweieler M, Joziasse J, Koelmans AA, 2005. Temporal variation of trace

metal geochemistry in floodplain lake sediment subject to dynamic hydrological conditions.

Environmental Pollution 137:281-294.

Van Griethuysen C, Meijboom EW, Koelmans AA, 2003. Spatial variation of metals and acid volatile

sulfide in floodplain lake sediment. Environmental Toxicology and Chemistry 22/3:457-465.


34

Van Griethuysen C, Meijboom EW, Koelmans AA, 2006. Temporal dynamics of AVS and SEM

insediment of shallow freshwater floodplain lakes. Applied Geochemistry 21:632-642.

Van Griethuysen C, Van Baren J, Peeters THM, Koelmans AA, 2004. Trace metal availability and

effects on benthonic community structure in floodplain lakes. Environmental Toxicology and Chemistry

23/3:668-681.

Vandegehuchte M., Roman Y., Nguyen L., Janssen C.R. and De Schamphelaere K. Toxicological

availability of nickel to the benthic oligochaete Lumbriculus variegates. Environ. Int., 33, 736-742.

Vijver M.G., Van Gestel C.A.M., Lanno R.P., Van Straalen N.M. and W.J.M. Peijenburg, 2004.

Critical review. Internal metal sequestration and its ecotoxicological relevance: a review. Environmental

Science and Technology, 38 (18), 4705-4712.

Wang W.X., Stupakoff I. and Fisher N.S., 1999. Bioavailability of dissolved and sediment-bound

metals to a marine deposit feeding polychaete. Mar. Ecol. Prog.Ser 178, 281-293.


35

ANNEX A: IMPORTANCE OF SPATIAL AND TEMPORAL VARIABILITY IN AVS

CONCENTRATIONS

Several studies reported on the dynamic behavior of AVS in natural systems. Besides the inherent

spatial variations observed between different sampling locations AVS concentrations differ with depth.

Most often the AVS concentration increases with increasing sediment depth (even over small sample

distances 0-10 cm,) and is linked to the redox gradient present in the sediment (Van Den Berg et

al, 1998 and Van Den Berg et al, 2001). In addition there seems to be a strong seasonal

component where AVS concentrations tend to be the higher at the end of the summer and during

fall and lower in winter and spring (Howard & Evans, 1993; Van Den Hoop et al, 1997;

Grabowski et al, 2001). Most often the above mentioned phenomena are strongly influenced by

dynamic behavior of the overlying water stream (Poot et al, 2007).

Temporal variability

Temporal variability has been addressed by several authors (Van den Berg et al. 1998; Van

Griethuysen et al., 2005, 2006). The results indicate that seasonal variability is closely linked to

microbial activity. Microbial activity depends on water temperature and higher temperatures in spring

and summer will result in increased microbial activity yielding a higher sulfate reduction rate. The net

result is hat AVS concentrations tend to be generally higher at the end of the summer and during

fall and lower in winter and spring when microbial activity is low. Care should be taken in

extrapolating the results of floodplain soils (e.g. Van Griethuysen et al, 2005/ Poot et al, 2007).

In fact some of the floodplain soils showed an opposite AVS seasonality because of preferential

inundation and concomitant AVS formation in winter as was observed by Poot et al, 2007.

The sampling strategy used for most of the AVS databases reported in this report took this

seasonality factor as much into account as possible. Samples were taken by preference in spring

time when AVS levels are expected to be the lowest.

Spatial variability

The importance of spatial heterogeneity of AVS and associated metal concentrations has been

recognized in risk assessment of trace metal polluted sediments. Two criteria are very important for

the spatial variability:


36

a) vertical variability (depth)

b) two dimensional horizontal variability (surface)

Vertical spatial variability

The observed vertical gradient in sediment AVS is mainly caused by the oxidation of AVS near the

sediment/overlying water interface. Another contributing factor is sediment bioturbation (De Witt,

1996, Peterson et al, 1996). Typically lower AVS concentrations are measured in the 0-5 cm

gradient due to the presence of an aerobic layer that is rather thin, typically on the order of only a

few millimeters to a few centimeters in thickness (Carlton and Klug 1990; Hesslein 1976; Statzner

et al. 1988 in DiToro et al. 1991). Any surficial layer metal sulfide that becomes dissolved in the

pore water, as a result of metal sulfide oxidation, will not simply build up in the pore water and

remain there. Rather it will be subject to diffusion from the pore water into the overlying water as it

is produced. Given that the aerobic layer is quit thin, this diffuse flux will tend to temper any

increase in pore water metal levels that occur as a result of the oxidation process.

It should be noted that AVS concentration in the top layer should not be set to zero by default

because of the assumption that they are aerobic. Studies showing depth profiles indeed still indicate

measurable amounts of AVS in the 0-5 cm layer. Example: between 1-2 cm AVS concentration

ranged from 0.7-6.1 µmol/g dw in a floodplain lake (Van Griethuysen et al, 2005). Similar results

were obtained for Swedish sediments (AVS concentrations between 0-2 cm ranged between 0.011

and 2.86 µmol/g dw (Wiklund and Sundelin, 2002). Hansen et al 1996 reported < 1 µmol/g for

sediment depths 0-1 cm, 1 to 8 µmol/g wt for 1-3 cm sediment depth. Finally Buykx et al,

1999 reported AVS concentrations between 0.8-4.2 µmol/g dry wt for a sediment depth of 0.5 cm.

Comparison of the 10th percentile of the Flanders database (0.77 µmol/g dry wt.; typical sampling

depth (0-20cm) with the above reported typical AVS concentrations in the 0-5 cm surface layers

clearly indicate that the 10th percentile of the bulk AVS concentration (i.e. 0.77 µmol/g dry wt) is

within the reported ranges for surface sediment layers and can be considered as such as a suitable

estimate of AVS concentrations in sediment surface layers.

In the most recent monitoring campaign (UK, Finland and Spain) sampling depth was restricted to

0-5 cm.

Horizontal spatial variability


37

Only a limited amount of papers actually investigated the spatial variability of AVS concentrations in

sediments. Most often the spatial variability described are for floodplains that are only occasionally

inundated during winter time and are exposed to more extreme conditions. Therefore it is difficult to

extrapolate these observations to river sediments which are constantly underwater.

Van Griethuysen et al. (2003) investigated the spatial variability 0-5 cm of the sediments at 43

locations in a floodplain lake in the Netherlands. SEM concentrations were more or less constant

(average 5.3 µmol/g ± 0.4) while AVS concentrations were highly variable (average 15.3 µmol/g

± 9 µmol/g) showing a strong spatial dependence due to differences in lake depth, total sulfur

pools and redox potential.

Van Griethuysen et al. (2004) also investigated AVS and ΣSEM concentrations for 10 floodplain

Dutch lakes in the month of September (good separation between nearby rivers and lake

sediments). Four locations per lake were sampled to account for spatial variability in lake depth,

grain size distribution, organic matter content and redox conditions. Since the inundations are often

occurring in the Netherlands, striking differences are observed in total trace metal distribution in

relation to OC and clay even between the same floodplain sediments. Moreover other criteria like

distance from nearby rivers, number of inundations per year, surface of lake, depth and water

transparency were taken into account in the interpretation of the results. The AVS values were in the

range 0.9-48.8 µmol/g, while the SEM between 0.2-14.0 µmol/g. Trace element content in

floodplains is positively correlated with the inundation frequency.

Poot et al. (2007) investigated the spatial variability of AVS within the framework of the EU

program Aquaterra. They collected sediments from six sampling sites in the river Dommel or its

tributaries and four floodplain soils adjacent to the river. At each site three sediment/soil cores within

1m2 (0-5 cm) were taken. Precision of the chemical analysis was evaluated by taking triplicate

measurements of SEM and AVS concentrations. Averaged relative standard deviation (RSD) of the

triplicate measurements on a homogenized sediment/soil sample (i.e. upper 5 cm of one core) was

9 % for AVS and 19 % for SEMCu. On request the author shared the raw data on the 3 individual

sediment cores taken within 1 m2 at one point in time. This allows us to actually evaluate the

spatial horizontal variability within a small surface area (1m2). The AVS and ΣSEM measurements

performed on each site showed that the standard deviations can be fairly large for both parameters

(Table A2).


38

Table A1: Relative standard deviation (RSD) with min-max indication between brackets (calculations

based on raw data provided by Poot et al, 2007)

Sediment type RSD AVS

(%)

RSD SEM (%) RSD OM (%)

River sediment 70.3

(3.3-173)

31.3

(1.6-115.6)

40.6

(2.9-126,6)

Flood plain

soil/sediment

44.3

(4.7-

97.9)

29.1

(5-81.8)

19

(2.5-46.4)

Relative Standard Deviation (RSD or coefficient of variation)) was for AVS on average 70.3 %

(3.3-173 %) for rivers and 44.3 % (4.7-97.9 %) for flood plains. For SEM RSD values of 31.1

% (1.6-115.6 %) were reported for rivers and 29.1 % (5-81.8 %) for SEM concentrations in

floodplains. Since the triplicate samples show a low RSD (<10% for AVS and 20% for ΣSEM),

the observed standard deviation is primarily caused by small scale spatial variation. The observed

spatial variability is of the same order of magnitude to the variability what can be observed for other

parameters frequently measured in sediments. For example, the RSD of the organic matter

measurements for river sediments have and average value of 40.6 % (2.9-126.6 %). Flood plain

soils/sediments showed lower RSD values with an average of 19 % (2.5-46.4). Birch et al

(2001) reported also RSD values in the same order of magnitude for metal concentrations in

aquatic sediments of dynamic environments.

Summary and conclusion

Both temporal and spatial variations are important to be considered when collecting SEM-AVS data.

AVS concentrations have the tendency to be lower in spring and winter then in summer. Furthermore

lower AVS levels are measured in the 0-5 cm sediment layer than in deeper layers.

The AVS database for Flanders contains a large number of sampling points (200), representative

for the Flemish river sediments. Therefore, the largescale geographical variability is well covered in

the sampling design. Furthermore samples were taken in spring season when it can be expected that

AVS concentrations are the lowest.

Both the one-sided Standard Two-sample t-Test, Welch modified Two-sample t-Test, Wilcoxon

rank-sum test and Two-sample Kolmogorov Smirnov Test were unable to detect statistical significant

difference between the 50th percentile values of the AVS databases for Finland, UK, Spain, Serbia

and the Netherlands when compared with the Flemish (Belgium) database. The data from Poot et


39

al, (2007) showed that the variability observed in AVS and SEM measurements is mainly related to

small-scale spatial variation and less with analytical errors. This small scale variability makes it

difficult to assess spatial variation between sites or temporal variation within a site.

Available AVS databases

In order to apply the SEM-AVS correction a good picture of the presence and geographical

distribution of AVS concentration in European sediments is needed. Knowledge on the occurrence of

AVS concentrations is, however, limited. An overview of the different databases per country is given

in Figure 1 as Box-Whisker plots with AVS data reported for other countries. A box and whisker

diagram, or boxplot, provides a graphical summary of a set of data based on the quartiles of that

data set:

The ‘box’, or rectangle, in Figure A1 contains 50% of the data, and the extremes of that box are

the 25th percentile and 75th percentile. Each ‘whisker’ represents the remaining 25% of the data

and the extremities of these whiskers are the minimum and maximum values of the data.

Figure A1: Overview Box-Whisker plots AVS data (µmol/g dry wt) for Belgium (Flanders, n =202 stations),

the Netherlands (n = 29 stations), Hungary (n = 9 stations), UK (n= 16 stations), Finland (= 25 stations),

0

0,01 0,1 1 10 100 1000

Concentration (µmol AVS/g dry wt)

Italy

Sweden

Serbia

UK

Spain

Hungary

Finland

Netherlands

Belgium


40

Spain (n = 20 stations), Serbia (n = 12 stations). For Italy (n = 4 stations) and Sweden (n = 4 stations)

only the individual sampling points are given.

The largest database available is for the Flemish region in Belgium. This database representative for

EU low midland rivers contains 200 sediments sampled over a depth of 0-10 cm. The 50th

percentile of the AVS distribution yields an AVS value of 8.7 µmol/g dry wt (Vangheluwe et al.

20051). The lowest AVS concentration in the Flanders dataset is 0.045 µmol/g dry wt. The 10th

percentile is 0.77 µmol/g dry wt. The latter value was suggested in the different ongoing metal risk

assessments to be used as a generic default correction value for low midland rivers (the

Netherlands, Germany and possibly Northern France) when site specific measurements are lacking.

The Netherlands (Van den Hoop et al, 1997, Van Den Berg et al, 1999, 2001)) has a database

of 29 sediments with a 50th percentile of 5.85 µmol/g dry wt and a 10th percentile of 1.2 µmol/g

dry wt. Limited data sets for Germany and Hungary are available. For Germany only data from one

site at the river Rhine (up and downstream local site) and the river Lippe (up and downstream

local site) and the artificial Schmallenberg pond (10 samples) are available but not considered as

relevant for a regional risk characterisation. AVS levels range from 0.1-16.2 µmol/g dry wt. with a

50th percentile of 5 µmol/g dry wt and a 10th percentile of 0.3 µmol/g dry wt. For Hungary

AVS data (n=9) are available for the river Tisza and Szamos. Most values were equal are above

the 0.77 µmol/g dry wt. benchmark. At three sites AVS concentrations were measured that were

lower: 0.1, 0.2 and 0.3 µmol/g dry wt., resulting in a 50th percentile of 0.5 µmol/g dry wt and

a 10th percentile of 0.18 µmol/g dry wt.

It has been suggested by various Member States that not enough data were available at present to

determine if the Flanders dataset is indeed representative of other EU regions. Therefore it was

agreed to embark on a multi-metallic Conclusion 1 program aimed at collecting AVS data for

countries that are not yet covered by the current dataset. As such 3 additional countries have been

sampled: Finland, UK and Spain (Vangheluwe et al, 2008)..

1 Vangheluwe, 2005. discussion document on the selection of an AVS default value representative

for the EU region, 21th November, 2005.


41

The sediment sampling program focused on sampling the 0-5 cm of the sediments and having a

representative picture of the regional AVS distribution. Although the intention was to sample in the

spring season (April-May) where AVS concentrations are expected to be the lowest, this was only

possible for Finland (most samples taken in May). In the UK, sampling was conducted in June-

September because flooding events in the period May-June preventing sampling. Spain could only

start the monitoring in Oktober. In Finland a total of 25 samples were taken (13 lakes, 12 rivers).

Analysis of the AVS concentrations gives a 10th percentile of 1 µmol/g dry wt and a 50th percentile

of 11 µmol/ g dry wt. The lowest concentration measured was 0.3 µmol/g dry wt. For the UK 16

sediments from 16 different rivers were sampled belonging to 8 EA regions (Anglian-, Midlands-,

Northeast-, Northwest-, Southwest-, Southern-, Thames and Wales region). Analysis of the AVS

concentrations gives a 10th percentile of 0.31 µmol/g dry wt and a 50th percentile of 7.95 µmol/ g

dry wt. The lowest concentration measured was 0.071 µmol/g dry wt. For Spain 20 samples of

the river Ebro were sampled. Analysis of the AVS concentrations gives a 10th percentile of 3.68

µmol/g dry wt and a 50th percentile of 13.5 µmol/ g dry wt. The lowest AVS concentration

measured was 1.7 µmol/g dry wt.

In addition other databases were recently identified. Prica et al (2007) analyzed SEM-AVS

concentrations in contaminated sediments collected from rivers in the Danube basin. Sediment was

analyzed in the following watercourses: the Begej, Tisa and Tamis rivers, the Danube-Tisa-Danube

(DTD) Canal, the Sava at Sabac and the Danube. Analysis of the AVS concentrations gives a 10th

percentile of 4.3 µmol/g dry wt and a 50th percentile of 7.43 µmol/ g dry wt. The lowest AVS

concentration measured was 3.1 µmol/g dry wt

Burton et al. (2007) investigated AVS concentrations for 84 sites in wadeable streams of 10

countries and nine ecoregions of Europe. The results showed AVS concentrations ranging from 0.004

µmol/g dry wt. to 44 µmol/g dry wt with an median value of 0.1 µmol/g dry wt (sample depth

0-5 cm) and an average value of 2.5 µmol/g dry wt. It should be noted that sediments in this

program were collected in head streams resulting in very low AVS and SEM levels.

weight of evidence analysis avs normalization nickel 28-06

Documents