Metal mobilization in the Gironde Estuary (France): the role of the soft mud layer in the maximum...

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Marine Chemistry 87 (2004) 1–13

Metal mobilization in the Gironde Estuary (France): the role of the

soft mud layer in the maximum turbidity zone

Sebastien Robert*, Gerard Blanc, Jorg Schafer, Gilbert Lavaux, Gwenael Abril

Traceurs Geochimiques et Mineralogiques, TGM-UMR-CNRS 5805, Universite Bordeaux I,

Avenue des Facultes, F-33405 Talence, France

Received 14 August 2002; received in revised form 14 May 2003; accepted 14 May 2003

Abstract

Vertical profiles of heavy metals (Cd, Cu, Cr, Fe, Mn, Ni and Pb) in the particulate and dissolved phases and redox sensitive

parameters (dissolved oxygen, nitrate, nitrite and ammonium) were analyzed for the first time in the continuum Maximum

Turbidity Zone (MTZ)–Fluid Mud (FM; 50 g l� 1 < suspended particulate matter (SPM) < 500 g l� 1)-consolidated sediment of

a macrotidal, highly turbid estuary: the Gironde. The results show that the fluid mud is a layer of intense metal mobilization due

to redox induced dissolution, releasing trace metals into the water column. This newly identified phenomenon creates a transient

situation characterized by the onset of diagenetic sequences in the FM, i.e. in the lower part of the water column. These

sequences overlie permanent diagenetic sequences in the consolidated sediment, similar to those typically observed in marine

and estuarine sediments. The discontinuity of dissolved Cd, Ni and Pb concentrations at the sediment surface indicates that Mn

reduction is faster in the FM than in the upper sediment, isolated from the oxic water column by the suboxic FM. Two separate

diagenetic signals are preserved, as the installation of the diagenetic sequence in the FM is faster than the molecular diffusion of

dissolved compounds through the FM–sediment interface. The diagenetic signal of trace metals (e.g. dissolved Cr maximum)

in the upper sediment layer near the FM–sediment interface is interpreted as a transient record of past hydrologic situations,

during which the absence of the FM layer permitted the installation of an oxic/suboxic front in the upper sediment. The

resuspension of the FM during the spring tide probably results in a new distribution of these elements in the estuary, where they

may be stabilized in the dissolved phase (e.g. by chloride complexes or dissolved organic compounds) or be adsorbed onto

reactive particles (e.g. freshly precipitated Mn oxyhydroxides).

D 2004 Elsevier B.V. All rights reserved.

Keywords: Diagenesis; Gironde Estuary; Sediment; Redox; Trace metals

1. Introduction

The Gironde Estuary in the southwest of France

drains one of Europe’s least industrialized regions.

0304-4203/03/$ - see front matter D 2004 Elsevier B.V. All rights reserve

doi:10.1016/S0304-4203(03)00088-4

* Corresponding author. Tel.: +33-5-4000-8834; fax: +33-5-

5684-0848.

E-mail address: [email protected] (S. Robert).

Nevertheless, this macrotidal estuary has been strong-

ly polluted by heavy metals, due to former mining and

ore treatment activities in the upper reaches of a

tributary (Lot River) catchment since the late 19th

century (Latouche, 1988; Jouanneau et al., 1990;

Lapaquellerie et al., 1996; Kraepiel et al., 1997; Blanc

et al., 1999; Boutier et al., 2000; Michel et al., 2000;

Schafer and Blanc, 2002; Schafer et al., 2002). Macro-

d.

S. Robert et al. / Marine Chemistry 87 (2004) 1–132

tidal estuaries are characterized by long residence

times of water and particles and the presence of a

maximum turbidity zone (MTZ). The Gironde Estuary

is characterized by very high turbidities, with concen-

trations of suspended particulate matter (SPM) ex-

ceeding 1 g l� 1 in surface water and several hundreds

of g l� 1 in bottom water. This MTZ is often located in

the low salinity region, but spreads up and down

estuary with seasonal river flow variations (Sottoli-

chio and Castaing, 1999). In the MTZ intense cycles

of sedimentation and erosion occur on different time

scales: tidal and seasonal. Erosion occurs at mid-ebb

and mid-flood, at high current velocities, whereas

tidal slacks are sedimentation periods. Fluid mud

(FM) within the MTZ (highly concentrated benthic

layers with SPM concentrations of several hundreds

of g l� 1) is generally found during neap tides and low

water discharge, when sedimentation exceeds erosion.

Fluid mud layers, separated from the turbid water

column by a lutocline (Abril et al., 1999), remain

partially stable for a few days and are most intensively

eroded during spring tides, especially in wet years

(Allen, 1972).

Physicochemical processes in estuaries have at-

tracted increasing interest in recent years because of

their consequences for the net fluxes of particulate and

dissolved materials from land to ocean (Comans and

Van Dijk, 1988; Apte and Batley, 1995; Tipping et al.,

1998; Michel et al., 2000). Maximum concentrations in

the dissolved phase typically observed for metals like

Cd, Ni and Zn in the low and mid-salinity ranges are

due to mobilization from estuarine particles (Turner

and Millward, 1993; Elbaz-Poulichet et al., 1996;

Kraepiel et al., 1997; Guieu et al., 1998). Considering

the water column to be oxic and the sediment influence

(diffusive fluxes) to be negligible, these studies con-

cluded that, after metal mobilization by estuarine

salinity effects, chloride complexation is the major

process stabilizing metals in the dissolved phase

(Boughriet et al., 1992; Ouddane et al., 1992). How-

ever, studies in marine and estuarine sediments have

shown a strong influence of redox conditions on trace

metal partitioning and speciation (e.g. Pedersen et al.,

1986; Finney et al., 1988; Shaw et al., 1990; Morford

and Emerson, 1999, Tseng et al., 2001). In particular,

Mn and Fe oxides are dissolved in suboxic conditions,

releasing adsorbed and co-precipitated trace metals

(Gobeil et al., 1981, 1987; Gaillard et al., 1986).

However, mid-estuarine maxima in trace metal con-

centrations observed in different estuaries are mainly

explained by desorption of labile metals from seaward

fluxing SPM, rather than to contributions of dissolved

metals from sediment pore waters (e.g. Chiffoleau et

al., 1994; Kraepiel et al., 1997; Liu et al., 1998).

Processes responsible for trace element mobilization

are of great interest, especially in estuaries suffering

from extensive metallic pollution, like the Gironde

Estuary.

In this study, for the first time, near-bottom profiles

of trace metals (Cd, Cu, Cr, Ni and Pb) in water, newly

deposited fluid mud and sediment are measured, to-

gether with major redox parameters, in the MTZ of the

Gironde Estuary. The aim of this study is to investigate

the behavior of trace metals exposed to extremely

turbid, suboxic conditions typically observed in the

transient fluid mud. We will examine whether these

conditions are sufficiently reducing and persistent to (i)

induce a diagenetic response of heavy metals in the

lower part of the water column and (ii) have an impact

on the diagenetic signal of metals in the consolidated

sediment. Emphasis will be laid on the distinction of

diagenetic signals in the temporary fluid mud and the

permanent consolidated sediment to understand wheth-

er these signals (i) are stationary or transient and (ii) are

controlled by the current temporary situation or reflect

past, different hydrologic situations.

2. Sampling and analysis

2.1. Sample collection

The SEDIGIR experiment was conducted on board

the RV Gwendrez in June 1996 during low river flow

(460 m3 s� 1), when there was a well developed MTZ,

located in the freshwater and low salinity regions

(0.5 < S < 6). During the cruise, a complete spring to

neap period of fluid mud deposition and resuspension

was characterized by nutrient measurements (Abril et

al., 1999, 2000) at an anchor location in the MTZ (KP

43, Fig. 1) over 14 days. During this period, three

vertical profiles representing the continuum MTZ–

FM-consolidated sediment were sampled. The first

profile was sampled during a low tide slack on June

19, in a situation representing mean tidal amplitude.

The second profile was sampled during a high tide

Fig. 1. Location of the sampling site in the Gironde Estuary. KP:

distance expressed in kilometers from the city of Bordeaux.

S. Robert et al. / Marine Che

slack on June 21, at neap tide. The third profile was

sampled during a low tide slack on June 26, at neap

tide. Salinity and SPM concentrations were measured

in all samples. Trace metal analyses were performed

for the profiles sampled the 19th and 21st of June.

Nutrient measurements were performed for profiles

sampled the 21st and 26th of June.

Two samplers were especially designed to simul-

taneously collect samples from the water column, the

fluid mud layers and the underlying consolidated

sediment. The first sampler (sampler A) enables the

retrieval of samples in the water column including the

fluid mud with a 20-cm resolution (Abril et al., 2000).

The second sampler (sampler B) collects samples

from the water column to the consolidated sediment.

Although the sampler permits a maximum resolution

of 5 cm in the water column, a lower resolution (20

cm) was chosen to obtain enough sample volume for

the different analyses. Sampling resolution in the

sediment was limited to 5 cm for the same reasons.

2.2. Analytical techniques

Samples used for trace metal measurements were

recovered in 200 ml acid cleaned, polypropylene

centrifuge vials. After centrifugation, the supernatant

was filtered using 0.4 Am Nuclepore filters to separate

the dissolved phase. Immediately after filtration, about

2/3 of the filtrate was acidified (1x, HNO3 65%

suprapure) and kept at 4 jC until analysis. Sampling

and conditioning were performed under an N2 atmo-

sphere to avoid sample oxidation.

Sulfate concentrations were obtained from non-

acidified filtered waters, stored in borosilicate flasks,

using ion chromatography (AFNOR, 1995, 1996;

Rouessac and Rouessac, 1997), with an accuracy of

better than 97%. Particulate samples obtained by

centrifugation were dried, ground and homogenized.

Sample aliquots (30 mg) were digested using 750

Al HCl (12 N), 2 ml HF (26 N) and 250 Al HNO3

(14 N) in closed Teflon bombs (Savilex). The bombs

were heated at 110 jC for 2 h and, after cooling, the

solution was evaporated to dryness and the residue

dissolved in 150 Al concentrated HNO3. Each sample

was left on a hotplate until the residue was completely

dissolved and then brought to 5 ml in a volumetric

flask using double-deionized water as reported in

Schafer et al. (2002). All acids were of suprapure

quality. Dissolved Cd, Cu, Ni and Pb were preconcen-

trated in a Class 100 clean room by chelation with

ammonium pyrrolidine dithiocarbamate/diethyl am-

monium diethyl dithiocarbamate (APCD/DDDC), ex-

traction into Freon and back extraction with HNO3

into deionized water (Danielsson et al., 1982).

The Cr, Fe and Mn concentrations were measured

by graphite furnace using Zeeman effect background

correction (Cr was preconcentrated in the furnace by

multiple injections, 5� 20 Al aliquots) or flame atomic

adsorption spectrometry (AAS), depending on the

metal concentration. Dissolved and particulate Cd,

Cu, Ni and Pb concentrations were measured using

AAS (graphite furnace or flame, depending on the

metal concentration) and inductively coupled plasma

mass spectrometry (ICP-MS). Comparison of the ana-

lytical data obtained by both techniques (i.e. ICP-MS

and AAS) typically showed differences in concentra-

tions of less than 10%. The most reliable analysis in

terms of accuracy, detection limit and reproducibility

was selected (Table 1). Quality control and determina-

mistry 87 (2004) 1–13 3

Table 1

Reliability of dissolved and particulate metal concentrations, comparison with international standards

Cadmium Copper Iron Manganese Chromium Nickel Lead

Accuracy (%) particulate 4 7 7 4 9 2 1

dissolved 3 2 7 0 0 7 8

Analytical reproducibility:

rsd (%)

4 3 4 2 2 1 1

Particulate phase detection

limit (Amol kg� 1)

0.025 0.5 500 250 1.5 1.5 0.25

Dissolved phase detection

limit (nM)

0.05 1 50 10 1.5 3 0.05

International standard sediment

concentration (Amol g� 1)

4.7E� 03 7.1 852 8.5 2.17 0.75 0.2

Uncertainty range

(95% confidence limits)

2.3E� 04 0.25 14.3 0.22 0.15 0.03 7.7E� 03

Measured concentration (Amol g� 1) 4.3E� 03 6.8 781 8 1.85 0.74 0.19

International standard water

concentration (nM)

0.12 25.5 1.79E+ 03 70.9 3.3 14.1 0.13

Uncertainty range (95%

confidence limits)

0.02 1.7 35.8 5.5 0.3 0.3 0.024

Measured concentration (nM) 0.14 23.3 1.95E+ 03 76.4 3.4 15.5 0.101

Accuracy is obtained by comparison with recommended values of the international standards: SLRS-3, SLEW-2, CRM 320, PACS-1.

Accuracy is calculated as follows: accuracy = 100� (measured concentration� certified concentration)/certified concentration.

Example for dissolved Cu: A= 100� ((25.5� 1.7)� 23.3)/25.5 = 2%.

Analytical reproducibility is rsd calculation obtained from 10 measurements of the same sample.

Detection limits are given as three times the rsd of analytical blanks.


tion of accuracywere performed by analysis of certified

international reference materials of river and estuarine

waters (SLRS-3, SLEW-2) and river and marine sedi-

ments (BCR-CRM 320, PACS-1). Analytical results

obtained for the particulate and dissolved reference

materials always differed by less than 10% from the

certified values and reproducibility was generally bet-

ter than 4% (rsd) for every metal examined (Table 1).

Fig. 2. Salinity and SPM profiles in water column–sediment system

during June 19, 21 and 26 tidal slacks.

3. Results

3.1. Fluid mud layers and major dissolved redox

species

On the basis of previous publications (Bassoulet and

Le Hir, 1998; Romana, 1999; Abril et al., 1999) and our

present results, the following terminology will be used

to characterize the different layers in the GirondeMTZ.

The layer from the surface to an SPM concentration of

50 g l� 1 is referred to as ‘‘Turbid Water’’ (TW). SPM

concentrations from 50 to 250 g l� 1 characterize the

‘‘Liquid Mud’’ (LM), a highly unstable layer being

reworked at the tidal time scale (Abril et al., 1999). The

oxic/anoxic interface is located in this layer at an SPM

concentration around 100 g l� 1. Below the oxic/anoxic

interface the zone of denitrification is located (Abril et

S. Robert et al. / Marine Chemistry 87 (2004) 1–13 5

al., 2000). The ‘‘Soft Mud’’ layer (SM), showing SPM

concentrations between 250 and 500 g l� 1, is generally

reworked at neap-spring time scale (Abril et al., 1999).

The SM is nitrate-free and ammonium-rich. The un-

derlying zone with SPM concentrations greater than

500 g l� 1 is considered as ‘‘consolidated sediment’’

that cannot be removed by bottom currents. The ‘‘Fluid

Mud’’ as usually defined in the literature includes the

LM and the SM.

The SPM–depth and salinity–depth profiles ob-

tained from the three sampled tidal slacks are presented

in Fig. 2. Two contrasting situations are observed. The

low tide situations on the 19th and 26th of June are

Fig. 3. SPM, salinity, dissolved oxygen, sulfate concentrations, dissolved M

developed benthic layers (—o— 19/06/1996; —w— 26/06/1996). Situati

similar, showing thick LM and SM layers (>1m; Fig. 2)

and quite homogeneous salinity distributions, whereas

the high tide situation on June 21 shows thinner LM

and SM layers (LM+ SM< 0.5 m) and decreasing

salinity with depth. This difference in SPM distribution

was attributed to local bathymetric effects rather than to

erosion and redeposition processes. Indeed, high fre-

quency SPM profiles (Abril et al., 1999) revealed a

relatively stable SM layer between these dates.

Concentrations of SPM, salinity and the redox status

of the different layers sampled are shown in Fig. 3. In

the water column, redox conditions confirm previous

descriptions by Abril et al. (1999, 2000). In the TW,

n and nutrients from two contrasting situations. Situation 1: highly

on 2: poorly developed benthic layers (—E— 21/06/1996).


nitrite, ammonium and dissolved Mn are below detec-

tion limits or unimportant, whereas dissolved oxygen

and nitrate concentrations are characteristic of oxic

waters (Fig. 3). The results obtained from the LM

samples of June 26 are typical for low discharge

situations with a highly turbid, well stratified MTZ,

i.e. dissolved oxygen and nitrate rapidly decrease with

depth, reaching undetectable levels near the LM–SM

interface (e.g. Abril et al., 1999; Tseng et al., 2001).

The SM shows a nitrite peak and an increase of am-

monium due to ammonification. In this layer on the

19th of June, high dissolved Mn concentrations indi-

Table 2A

Particulate phase concentrations of Cd, Cu, Fe, Mn, Cr, Ni, Pb

Sample Deposited sediment

distance (m)

Cd

(nmol g� 1)

Cu

(Amol g� 1)

F

(A

19 June A1 2 6.2 0.59 85

A2 1.8 5.2 0.58 80

A3 1.6 5.1 0.56 80

A4 1.4 6.7 0.61 82

A5 1.2 4.2 0.56 80

A6 1 3.7 0.35 59

A7 0.8 n.m. 0.43 63

A8 0.65 4.2 0.39 62

A9 0.6 5.4 0.5 80

A10 0.45 4.3 0.36 58

A11 0.4 5.5 0.25 58

A12 0 6.3 0.53 64

A13 � 0.05 5.3 0.32 48

A14 � 0.1 5.2 0.39 64

A15 � 0.15 5.0 0.35 58

A16 � 0.2 5.1 0.34 62

A17 � 0.25 6.9 0.52 62

21 June B1 2.04 8.3 0.54 71

B2 1.84 5.3 0.5 69

B3 1.64 4.8 0.51 70

B4 1.44 4.5 0.55 79

B5 1.24 4.9 0.57 75

B6 1.04 4.0 0.43 66

B7 0.84 4.0 0.54 76

B8 0.64 5.9 0.56 80

B9 0.54 4.4 0.55 82

B10 0.44 3.3 0.23 56

B11 0.24 4.5 0.5 71

B12 0.09 4.3 0.54 76

B13 0.04 4.1 0.33 54

B14 � 0.01 3.0 0.17 37

B15 � 0.06 1.4 0.00 6

B16 � 0.11 1.4 0.00 16

B17 � 0.16 3.2 0.21 35

B18 � 0.21 3.6 0.26 41

n.m.: not measured.

cate reduction of Mn oxides. Under similar conditions,

dissolved Mn profiles obtained by Abril et al. (1999)

showed that dissolved Mn concentrations in an anoxic

fluid mud may reach values of up to 50 Amol l� 1.

Dissolved Fe concentrations do not show any impor-

tant increase, suggesting the absence of significant

reduction of Fe oxyhydroxides in the SM. The meas-

urements show no evidence for sulfate reduction in the

SM layer. In the consolidated sediment, dissolved Mn

and Fe show distinct peaks due to intensive dissolution

of Mn and Fe oxides. In the deeper consolidated

sediment sampled on June 19, sulfate concentration

e

mol g� 1)

Mn

(Amol g� 1)

Cr

(Amol g� 1)

Ni

(Amol g� 1)

Pb

(Amol g� 1)

3 15.4 2.3 0.81 0.33

7 16.1 2.15 0.75 0.32

1 15.7 2.17 0.77 0.33

7 16.8 2.22 0.77 0.34

7 12.6 1.97 0.71 0.27

4 9.2 1.75 0.58 0.2

2 13.1 1.91 0.63 0.22

6 11.8 1.9 0.64 0.22

1 15.0 2.45 0.83 0.34

1 11.2 1.83 0.6 0.21

1 14.5 1.62 0.54 0.2

5 16.8 1.96 0.62 0.27

4 13.9 1.29 0.42 0.19

5 16.1 1.79 0.6 0.25

1 14.2 1.65 0.58 0.22

0 16.0 1.79 0.6 0.27

6 14.8 1.87 0.64 0.28

7 12.8 2.16 0.75 0.28

1 11.2 2.08 0.74 0.27

4 11.2 2.3 n.m. 0.28

4 15.6 2.38 0.82 0.35

5 14.7 2.27 0.8 0.32

5 11.8 2.03 0.71 0.25

8 13.3 2.26 0.82 0.31

7 16.8 2.15 0.77 0.34

7 17.6 2.1 0.8 0.35

8 10.9 1.48 0.52 0.19

7 14.3 1.82 0.66 0.28

8 16.3 2.92 1.12 0.36

8 12.2 1.8 0.62 0.23

3 8.9 4.63 1.98 0.19

3 2.6 2.56 1.15 0.1

0 4.1 0.48 0.17 0.12

4 7.4 1.02 0.33 0.18

2 9.5 1.1 0.36 0.19

Table 2B

Dissolved phase concentrations of Cd, Cu, Fe, Mn, Cr, Ni, Pb

Sample Deposited sediment

distance (m)

Cd

(nM)

Cu

(nM)

Fe

(AM)

Mn

(AM)

Cr

(nM)

Ni

(nM)

Pb

(nM)

19 June A1 2 0.23 29.0 0.46 < 0.01 n.m. 3.3 0.32

A2 1.8 0.59 33.4 1.04 < 0.01 n.m. 5.5 0.23

A3 1.6 0.29 42.4 0.9 < 0.01 n.m. 5.2 0.6

A4 1.4 0.61 29.6 1.3 < 0.01 n.m. n.m. n.m.

A5 1.2 0.27 32.9 0.68 < 0.01 9.2 5.6 0.47

A6 1 0.32 31.6 1.43 0.2 6.9 6.1 0.84

A7 0.8 < 0.17 50.7 0.78 0.33 4.6 16.2 n.m.

A8 0.65 0.7 39.6 0.39 < 0.01 8.2 6.6 1.22

A9 0.6 0.73 42.4 1.03 0.03 5.8 11.5 0.94

A10 0.45 0.47 51.1 1.1 < 0.01 3.6 9.2 0.98

A11 0.4 1.65 51.1 0.2 1.23 5.1 11.6 2.87

A12 0 1.72 107 n.m. 7.1 3.1 22.6 1.97

A13 � 0.05 1.67 84.3 0.68 45.9 n.m. n.m. n.m.

A14 � 0.1 < 0.17 36.0 0.44 70.7 46.3 n.m. 0.01

A15 � 0.15 < 0.17 16.4 0.64 105 11.4 n.m. 0.04

A16 � 0.2 0.19 23.5 6.43 256 10.0 35.9 0.56

A17 � 0.25 0.48 16.3 n.m. 329 9.4 37.1 3.69

21 June B1 2.04 0.97 24.4 0.15 < 0.01 9.1 6.7 0.76

B2 1.84 0.93 24.8 0.48 < 0.01 8.9 6.2 0.44

B3 1.64 1.01 25.6 0.44 < 0.01 6.3 8.4 0.28

B4 1.44 0.84 25.2 0.61 < 0.01 5.9 7.8 0.33

B5 1.24 0.74 76.7 0.18 < 0.01 5.9 7.6 0.27

B6 1.04 0.7 26.6 0.23 < 0.01 5.6 6.4 0.47

B7 0.84 0.64 27.5 0.58 < 0.01 5.6 8.4 0.33

B8 0.64 0.53 24.0 0.53 < 0.01 4.7 6.6 0.28

B9 0.54 0.61 34.9 0.15 < 0.01 4.6 6.8 0.42

B10 0.44 0.55 24.3 0.44 < 0.01 5.0 6.0 0.43

B11 0.24 0.5 28.8 0.05 < 0.01 4.5 6.5 0.45

B12 0.09 0.64 29.9 0.15 < 0.01 4.5 6.5 0.42

B13 0.04 1.49 76.3 0.59 3.16 6.3 26.8 1.51

B14 � 0.01 < 0.17 584 0.35 5.82 1.9 7.4 0.52

B15 � 0.06 < 0.17 386 0.59 0.2 27.2 0.8 0.2

B16 � 0.11 < 0.17 121 0.44 < 0.01 16.0 0.8 0.1

B17 � 0.16 1.15 n.m. 0.35 8.41 11.7 6.9 0.25

B18 � 0.21 1.34 74.3 1.29 67.4 11.0 41.2 3.85

n.m.: not measured.

Table 3

Comparison of dissolved and particulate metal concentrations (Cd, Cu, Fe, Mn, Cr, Ni, Pb) of 8 and 10 m deep oxic samples with concentrations

of surface samples at comparable salinities

Cd Cu Fe Mn Cr Ni Pb

p d p d p d p d p d p d p d

A1: 8 m deep sample, S= 2 6.2E� 03 0.2 0.6 29.1 853 470 15.5 < 10 2.3 9.2 0.8 3.3 0.3 0.3

Surface S= 3.2 (Kraepiel et al., 1997) 4.5E� 03 0.5 0.5 15.6 938 24 11.5 28 0.9 7.1 0.2 0.2

B1: 10 m deep sample, S= 5.9 8.3E� 03 1 0.5 24.4 717 150 12.8 < 10 2.2 9.1 0.8 6.7 0.3 0.8

Surface S= 6 (Kraepiel et al., 1997) 5.1E� 03 1 0.6 18.7 906 21 13.7 19.5 0.9 8.1 0.3 0.3

p means particulate phase concentration in Amol g� 1, d dissolved phase concentration in nM.

Surface data from Kraepiel et al. (1997).



decreases despite increasing salinity, which indicates

the onset of sulfate reduction (a brown–black interface

was visible in the sediment at about 10 cm depth).

3.2. Particulate and dissolved metal concentrations

Particulate and dissolved metal concentrations

measured in the profiles are reported in Tables 2A

and 2B. As particulate metal concentrations are at

least 70 times higher than dissolved concentrations,

Fig. 4. Dissolved concentrations of Mn, Fe, Cr, Cu, Cd, Pb and Ni in the

developed benthic layers (—o— 19/06/1996). Situation 2: poorly develo

the variations in particulate metal content cannot be

attributed to metal mobilization, but rather to varia-

tions in grain-size distribution within the different

SPM layers (Jouanneau and Latouche, 1981).

Particulate and dissolvedmetal concentrations in the

upper, oxic part of the profiles (sampleA1 andB1 at 8m

and 10 m depth, respectively) were compared to those

reported by Kraepiel et al. (1997) for the surface water

at the same salinity in February 1994 (Table 3). The two

data sets are in the same range, except for dissolved Fe

Gironde Estuary from two contrasting situations. Situation 1: highly

ped benthic layers (—E— 21/06/1996).


concentrations that are more than one order of magni-

tude higher than the values reported by Kraepiel et al.

(1997). On the other hand, they are an order of mag-

nitude lower than those reported by Tseng et al. (2001)

for the MTZ in the Gironde. However, our data (150

and 470 nmol l� 1 at salinities of 6 and 3, respectively)

are close to mean dissolved Fe concentrations usually

measured in the Gironde Estuary (Boust et al., 1999)

and to values reported for the Seine Estuary at compa-

rable salinities (around 360 and 540 nmol l� 1 at

salinities of 6 and 3, respectively; Boust et al., 1999).

Dissolved metal profiles in the TW–LM–SM–

sediment continuum are presented in Fig. 4. In the two

profiles sampled on June 19 and 21, dissolved metals

show similar distributions in the different density

layers. Dissolved Mn, Cd, Ni and Pb concentrations

show a two-peak distribution in the SM and sediment.

Their concentrations show low variations in the TW

and the LM layers, but significantly increase in the

SM (Fig. 3 and Table 2B). Just below the SM–

sediment interface, the concentrations of Cd, Pb and

Ni in the pore water decrease to minimum values,

even lower than those in the TW and LM. Deeper in

the sediment, dissolved Mn, Cd, Ni and Pb concen-

trations increase and reach maximum values for Mn,

Ni and Pb (Fig. 4). In contrast, dissolved Cr and Cu

concentrations are nearly constant and low in TW, LM

and SM, with one very distinct peak near the SM–

sediment interface and decreasing values towards the

bottom of the core (Fig. 4). However, the Cu peak is

located at the interface, whereas maximum Cr con-

centrations in pore water occur at 10 and 6 cm below

the SM–sediment interface, on June 19 and 21,

respectively (Table 2B). Dissolved Fe concentrations

are constant in the FM and the upper sediment layer.

Only one sample at the bottom of the core of June 19

shows an increased dissolved Fe concentration, below

the zone of dissolved Mn increase (Fig. 4).

4. Discussion

4.1. Diagenetic response of metals in the SM

The data suggest that the settlement of the SM

layer caused hypoxia in the lower water column

creating a new, temporary, diagenetic sequence above

the classic diagenetic sequence in the consolidated

sediment. Permanent diagenetic sequences have been

reported by Shaw et al. (1990) in Californian border-

land sediment, by Chaillou et al. (2002) in Bay of

Biscay sediments and by Widerlund and Ingri (1995)

in Kalix Estuary sediments. In the deeper zone of the

consolidated sediment, reduction of Mn and Fe oxy-

hydroxides is observed. These diagenetic processes

typically release various trace metals such as Cd, Ni,

Pb, Zn into the dissolved phase (Shaw et al., 1990;

Morford and Emerson, 1999) due to destruction of

sorption sites by dissolution of Mn and Fe oxyhydr-

oxides (Guieu et al., 1998). Although anaerobic

sulfate reduction occurs in the deepest sample of the

studied core, the data do not support conclusions on

potential metal sulfide precipitation at this depth.

According to the classic scheme of early diagenesis

in sediments, the major diagenetic processes in the FM

(i.e. LM+SM) can be described as follows: aerobic

respiration and denitrification mainly take place in the

LM, whereas the SM is characterized by the onset of

Mn oxide reduction. Manganese reduction in the SM is

a typical phenomenon reported for different neap tide

situations in theGironde Estuary (e.g. Abril et al., 1999;

Tseng et al., 2001). No significant increase in dissolved

Fe concentration is observed in the SM, which may be

explained by the fact that early diagenetic dissolution of

Mn oxides generally precedes that of Fe oxides (Bur-

dige, 1993). Assuming that redox conditions in the SM

are not sufficiently reducing or persistent to result in Fe

dissolution, the observed release of Cd, Ni and Pb is

probably linked to Mn reduction.

At the SM–sediment interface, two zones of in-

tense reduction and trace metal release are clearly

separated by very low dissolved Cd, Ni and Pb

concentrations (Fig. 4). The existence of two super-

imposed diagenetic sequences separated by a discon-

tinuity in dissolved metal concentrations at the SM–

sediment interface suggests that intense Mn reduction

is installed very rapidly in the SM. In contrast, there is

no evidence for Mn reduction in the upper layer of the

consolidated sediment, although isolated from the

oxic water column by the suboxic SM. Two separate

diagenetic signals are preserved, as the installation of

the diagenetic sequence in the SM is faster than the

molecular diffusion of dissolved compounds through

the SM–sediment interface.

In contrast to Cd, Ni and Pb, dissolved Cr and Cu

concentrations do not increase in the SM, but in the


upper sediment layer. Near surface peaks of Cu and Cr

at the oxic/suboxic front have been reported for sedi-

ments from the Californian borderland and the Iberian

coast (Shaw et al., 1990; Blasco et al., 2000). Shaw et

al. (1990) assumed that these metals were released from

organic matter by oxidative respiration near the sedi-

ment surface and further re-adsorbed onto particles

under reduced conditions deeper in the sediment. The

dissolved Cr and Cu distributions in the Gironde

sediment seem to follow the scheme described by

Shaw et al. (1990). As Cr(VI) is thermodynamically

favored in the dissolved phase (Abu-Saba et al., 2000)

and Cr(III) is very particle active (except for Cr(III)

stabilized by colloidal organic matter; Abu-Saba and

Flegal, 1995), Cr in pore water should be present in

its hexavalent species. In the zone of Fe reduction (at

ca. 20 cm depth), soluble Cr(VI) is reduced to particle

active Cr(III) that is rapidly removed from the pore

water by adsorption, probably by insoluble large

molecular humic matter associated with filter retained

particles (Guo and Santschi, 1997; Sedlak and Chan,

1997; Shaw et al., 1990).

Given the slow oxidation kinetics of Cr (45 to 90

days; Abu-Saba and Flegal, 1995), the elevated dis-

solved Cr concentrations in the upper sediment sug-

gest long oxygenation periods due to downward

diffusion of oxygen from the oxic water column into

the sediment. The Cr peak indicates the resulting oxic/

suboxic front and, thus, oxygen penetration depth

(Shaw et al., 1990). However, during the sampling

period the upper layer of the consolidated sediment

was isolated from the oxic water column by the

suboxic SM. Consequently, the diagenetic signal

observed near the FM–sediment interface is inter-

preted as a transient record of past hydrologic periods

during which the absence of the FM layer permitted

oxygen penetration into the sediment. Assuming the

absence of SM during the previous spring tide period,

oxygen distribution was quite homogeneous within

the entire water column. Since the SM had settled a

few days before sampling, our chemical recording

must be considered as a snapshot of a transient

situation (neap tide and tidal slack).

4.2. Mobilization and recycling of metals in the SM

The Gironde Fluid Mud, and in particular the SM

layer, appears to be a site of intense mobilization of

metals. In Fig. 5, we compare the distribution of

dissolved metals as a function of salinity (0–10 range)

for all samples (TW, LM and SM, excluding consol-

idated sediment) to results from Kraepiel et al. (1997).

In the 2–6 salinity range, our data for dissolved Ni, Pb

and Cd in the TW and the LM are similar to the values

reported by Kraepiel et al. (1997) for surface water in

February 1994. Although some concentration differ-

ences occur over the salinity gradient (probably due to

hydrological or seasonal differences between sampling

campaigns), the relation between metal concentration

and salinity is comparable for the two data sets.

Nevertheless, around salinity 2, dissolved Ni, Pb,

Cu and Cd concentrations in samples derived from

suboxic, nitrate-free soft mud (full circles; Fig. 5) differ

strongly from the general relationship between con-

centration and salinity. This difference highlights the

additional heavy metal release (Ni, Pb and Cd) by Mn

reduction due to intense degradation of organic matter

(Cu mobilization), independent from salinity effects.

The dissolved Cd maximum in the low salinity

region of the Gironde cannot result only from Cd

desorption from surface water particles (Kraepiel et

al., 1997). However, this Cd addition can probably be

explained by the highly heterogeneous vertical SPM

distribution in the MTZ. Indeed, the near bottom

layers of the MTZ represent a major pool of particu-

late metals in the estuary. Although the SM is limited

to the upper estuary and to low discharge periods and

neap tides, it may contain up to 70% of the particles

concentrated in the MTZ (4–5 106 t; Jouanneau and

Latouche, 1981). Resuspension of the SM layer by

spring tides could lead to additional metal release to

the dissolved phase due to desorption and stabilization

of dissolved metals by salinity effects (e.g. chloride

complexation of Cd; Comans and Van Dijk, 1988;

Thouvenin et al., 1997). Other metals such as Cu

could be stabilized by dissolved hydroxides or humic

complexes (e.g. Mantoura et al., 1978).

Redox oscillations due to repetitive cycles of SM

deposition and resuspension could change surface

properties of the particles (e.g. by repetitive dissolu-

tion and precipitation of Mn oxyhydroxides). Conse-

quently, successive redox cycles could enhance trace

metal cycling in estuaries by processes (i) releasing

metals that are not directly available for desorption

and/or (ii) generating particulate transport phases (e.g.

amorphous oxyhydroxides) with increased sorption/

Fig. 5. Dissolved metal concentrations versus salinity in the Gironde Estuary.


desorption capacity. These particulate transport phases

could be adsorbed onto reactive particles of low

settling velocities, which have flushing times similar

to water and solutes (Alber, 2000), and thus reduce the

mean residence time of metals in the estuary.

5. Conclusions

For the first time, vertical profiles were simulta-

neously sampled in the water column (including

freshly formed fluid mud) and in the consolidated

sediment of a macrotidal highly turbid estuary. The

vertical distributions of major redox parameters and

heavy metals in the water column indicate that the soft

mud is a layer of intense metal mobilization due to

redox induced dissolution. This newly identified phe-

nomenon creates a transient situation characterized by

the onset of diagenetic sequences in the SM, i.e. in the

lower part of the water column, overlying permanent

diagenetic sequences in the consolidated sediment.

The discontinuity of dissolved Cd, Ni and Pb con-

centrations at the sediment surface indicates that Mn

reduction is faster in the SM than in the upper

sediment, where diffusive exchange is limited.

The resuspension of the SM during spring tide

probably results in a new distribution of these ele-

ments in the estuary. Elements released by redox


processes in the SM may be stabilized in the dissolved

phase (e.g. by chloride complexes or dissolved organ-

ic compounds) or be adsorbed onto reactive particles

(e.g. freshly precipitated Mn oxyhydroxides). Further

research is needed to determine whether these par-

ticles will form permanently suspended particles,

which are rapidly transported towards the ocean, or

if they will be recycled in the MTZ.

Acknowledgements

This study is the scientific contribution number of

D.G.O. UMR CNRS 5805 EPOC, Bordeaux 1

University. The authors wish to thank the staff of the

‘‘Cote d’Aquitaine’’ research vessel. Y. Lapaquellerie

and N. Maillet are also gratefully acknowledged for

their important technical support. We thank M. Fell for

revision of the English grammar and syntax. This work

is part of the GIS ECOBAG program and was

financially supported by the Agence de l’Eau Adour-

Garonne, the French Ministry of Environment (LI-

TEAU program) and by a Marie Curie fellowship of

the European Community programme ‘‘Energy, Envi-

ronment and Sustainable Development’’ under con-

tract number EVK1-CT-2000-5003.

Associate editor: Dr. David Turner.

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