Factors affecting metal concentrations in reed plants (Phragmites australis) of intertidal marshes in the Scheldt estuary

e c o l o g i c a l e n g i n e e r i n g 3 5 ( 2 0 0 9 ) 310–318

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Factors affecting metal concentrations in reed plants (Phragmites australis) of intertidal marshes in the Scheldt estuary G. Du Laing ∗ , A.M.K. Van de Moortel, W. Moors, P. De Grauwe, E. Meers, F.M.G. Tack, M.G. Verloo Laboratory of Analytical Chemistry and Applied Ecochemistry, Ghent University, Faculty of Bioscience Engineering, Coupure Links 653, B-9000 Ghent, Belgium

a r t i c l e

i n f o

a b s t r a c t

Article history:

We aimed to identify the environmental factors which significantly affect metal uptake by

Received 24 October 2007

reed plants in the intertidal marshes along the river Scheldt. Transfer coefficients, defined

Received in revised form

as the ratio of metal concentrations in reed stems to the metal contents in specific sediment

7 January 2008

fractions (i.e. the exchangeable Cd and Zn fraction and total Cr, Cu, Ni and Pb content), were

Accepted 9 January 2008

calculated for each sampling site. They were inversely related to the sediment clay and/or organic matter content. Metal mobility and thus plant availability is higher in sediments with a lower clay or organic matter content. Moreover, the plants might actively accumu-

Keywords:

late in particular essential elements when concentrations in the sediments are rather low,

Phragmites australis

which is the case in sediments low in clay and organic matter contents. Finally, more sandy

Reed

sediments are expected to be susceptible to occasional oxidation of sulphides, which leads

Metal

to an increased metal availability. A higher salinity promoted the uptake of Cu, Cr and Zn. © 2008 Elsevier B.V. All rights reserved.

Wetland Plant uptake Schelde Scheldt Sediment Salinity Bioavailability

1.

Introduction

While many engineering studies on treatment wetlands adopt a black box approach analysing levels in the influent and effluent, more should be known about the patterns and processes of metal uptake, distribution and removal by different species of wetland plants. The extent of uptake and how metals are distributed within wetland plants can have important effects on the residence time of metals in plants and in wetlands, and

∗

Corresponding author. Tel.: +32 9 2645995; fax: +32 9 2646232. E-mail address: [email protected] (G. Du Laing). 0925-8574/$ – see front matter © 2008 Elsevier B.V. All rights reserved. doi:10.1016/j.ecoleng.2008.01.002

the potential release of metals. This information is needed to better understand these systems and to assure that the wetlands do not themselves eventually become sources of metal contamination to surrounding areas, as could occur during litter decomposition (Weis and Weis, 2004; Du Laing et al., 2006). Various works have examined metal uptake and distribution in common reed plants (Phragmites australis (Cav.) Trin. ex Steud.) (e.g. Larsen and Schierup, 1981; Schierup and Larsen,

e c o l o g i c a l e n g i n e e r i n g 3 5 ( 2 0 0 9 ) 310–318

1981a,b; Gries and Garbe, 1989; Peverly et al., 1995; Keller et al., 1998; Windham et al., 2003; Lesage et al., 2007a,b). In addition, Vymazal et al. (2007) recently reviewed trace metal concentrations in reed plants growing in natural and constructed wetlands. Most studies report the highest concentrations of most metals in the roots, while leaf tissue has the second highest concentrations followed by stems and rhizomes. Schierup and Larsen (1981a,b) and Peverly et al. (1995), however, found high concentrations of Zn aboveground, while other metals were retained in the roots and rhizomes. Restriction of upward metal movement into shoots can be a tolerance mechanism of wetland plants, next to restricting uptake, sequestering metals in tissues or cellular compartments (e.g. central vacuoles) that are insensitive to them and translocation of excessive metals into old leaves shortly before their shedding (McCabe et al., 2001; Windham et al., 2003; Weis and Weis, 2004). The protective mechanisms against metal toxicity differ between plant species (Fediuc and Erdei, 2002). A small fraction of metals is released in the environment during the growing season through living aerial leaf tissue. Although the exact nature of this release is uncertain, it may be the result of leaching from leaf surfaces accompanying transpirational water loss. This loss may be similar to the release of nutrients from leaves of wetland plants (Burke et al., 2000). Ye et al. (1997) compared different populations of Phragmites australis, one from a contaminated mine site and three from clean sites, and found similar Zn, Pb and Cd uptake. The metal-contaminated population thus did not seem to modify its uptake or distribution of metals as a response to the contaminated environment. Schierup and Larsen (1981a,b) investigated the uptake of heavy metals by reed plants in a polluted and a non-polluted Danish lake. Metal contents in reed plants of the non-polluted lake were higher than those in the polluted lake. They attributed this to differences in factors such as pH and redox potential, which are more important in determining metal uptake than total pollutant levels in sediments and surface waters. Ye et al. (1998) observed smaller biomass production and greater metal accumulation (especially Zn) in Phragmites australis seedlings grown for 90 days on a metal-contaminated sediment under flooded conditions compared to dry conditions. The redox potential facilitated Fe and Mn oxide reduction, but not sulphide formation. This suggested that metal availability to the plants was enhanced due to their release from Fe/Mn oxides that were being dissolved in reducing conditions. Lehtonen (1989) compared the metal accumulation by a series of aquatic macrophytes growing in an acidified lake and a lake with a fairly good buffer capacity. The accumulation in Phragmites australis was found to be little influenced by the lake acidity, in contrast to the species Nuphar lutea and mosses growing at the bottom of the lakes, who showed a major response. This was attributed to the fact that not all sediment layers from which reed plants obtain their nutrients are affected by acidification. Phragmites indeed takes up nutrients from sediments alone, whereas Nuphar lutea obtains nutrients from both water and sediments and the mosses only from the water. Windham et al. (2001) observed increased Pb uptake and translocation to aboveground parts of reed plants when adding Pb in the form of Pb-acetate to sediments. Keller et al. (1998), however, found no correlations between concentrations of extractable metals from marsh sediments and concentrations in roots of reed

311

plants, and suggested that factors other than labile concentrations in sediments control metal uptake and storage in roots. Similarly, weak relationships between metal concentrations in plants and phytoavailable stocks in the soil were previously observed for other plant species in other floodplains (e.g. Overesch et al., 2007). We aimed to check whether metal concentrations in reed plants along the river Scheldt agree with those previously reported in literature for other sampling sites. Moreover, we searched for relationships between sediment properties and metal concentrations in reed stems, to identify the environmental factors that are most dominant in determining metal uptake by these plants. Reed stems were selected to study the factors affecting metal uptake rather than roots or leafs. Metal concentrations analysed in roots and rhizomes may reflect some proportion of metals that are merely adsorbed onto the root surface rather than within the root tissue (Hall and Pulliam, 1995; Weis and Weis, 2004). Leaves might be more susceptible to atmospheric deposition compared to stems, due to their larger external surface area. Moreover, Windham et al. (2003) (in: Weis and Weis, 2004) reported that only 4–20% of the aboveground metals in reed plants were found in leaf tissue because stem biomass production is much higher.

2.

Materials and methods

The study was carried out in the part of the Scheldt estuary that is subjected to tidal influences, downstream of the city of Ghent. All 26 study sites are tidal marshes vegetated by a monospecific stand of common reed, Phragmites australis (Cav.) Trin. ex Steud. Most of them are inundated at springtide only. Three reed plants were randomly harvested at each sampling site in August within a distance of approximately 5 m from each other. The reed plant samples in the current study were separated into rhizomes, roots, leaves and stems and plumes. Leaf sheaths were removed and all leaves of each plant were pooled. The samples were washed with deionised water, dried at 70 ◦ C for 48 h, ground in a hammer-cross beater mill and homogenized. Plant samples were analysed for metal contents by weighing 1 g in beakers. Five milliliters of ultra pure HNO3 were added, after which the suspension was heated on a hot plate at 130 ◦ C during 1 h. After heating, a total of 4 ml hydrogen peroxide was added to each sample and the suspension was heated for another 10 min. Finally, the suspensions were filtered in volumetric flasks using filter papers (S&S blue ribbon, Schleicher & Schuell) and diluted to 50 ml using 1% HNO3 . The Cd, Cr, Cu, Ni, Pb and Zn concentrations were analysed using ICP-OES (Varian Vista MPX, Varian, Palo Alto, CA) and/or GF-AAS (Varian SpectraAA800/GTA-100, Varian, Palo Alto, CA). As it is very difficult to completely digest reed plant material preceding the analyses, destruction methods were compared and quality control results were described by Du Laing et al. (2003). The upper 0–20 cm sediment layer was sampled and analysed according to procedures given in Du Laing et al. (2007c). Summary statistics of metal concentrations in the sediments and general sediment properties are given in Table 1 (Du Laing et al., 2007c).

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Table 1 – Summary statistics of metal concentrations and sediment properties in the upper sediment layer of 26 intertidal marshes along the river Scheldt (n = 78; 3 samples taken at each sampling site) (Du Laing et al., 2007c) Mean −1

Cd (mg kg DM) Cr (mg kg−1 DM) Cu (mg kg−1 DM) Ni (mg kg−1 DM) Pb (mg kg−1 DM) Zn (mg kg−1 DM) % Clay % Silt % Sand % OM % CaCO3 Cl (mg kg−1 DM) pH EC (␮S cm−1 )

3.

7.6 134 97 37 143 595 32 43 25 11.6 7.6 393 7.6 730

Minimum 0.5 26 4 8 13 37 5 <1 <1 1.1 2.5 <2 7.2 130

10th percentile 1.7 45 19 13 37 132 13 18 1 2.9 4.9 <2 7.3 250

Results and discussion

Metal concentrations in the leaves and stems of Phragmites australis at the intertidal marshes along the river Scheldt are presented in Fig. 1, as a function of the distance to the river mouth, which is situated near Vlissingen (The Netherlands). Metal concentrations in leaves and stems of Phragmites australis found in this study were compared with other studies (Tables 2 and 3). Differences between studies might be

Median 7.5 147 96 38 146 650 35 51 9 12.2 7.1 21 7.5 530

90th percentile 13.0 195 166 58 218 868 47 59 71 19.9 10.5 1754 8.1 1790

Maximum 21.9 244 513 71 459 1501 48 66 95 23.3 13.4 3993 8.7 3720

related to pollution levels and physico-chemical sediment, water or sediments characteristics at the sampling sites. However, observed metal concentrations might also depend on the time of sampling. Most authors sampled reed plants for metal analysis during summer months, which renders the data comparable to our study, whereas some authors studied metal fluctuations in reed plants over the complete growing season (e.g. Larsen and Schierup, 1981; Scholes et al., 1999; Weis et al., 2003). Weis and Weis (2004) found that it is very difficult to generalize about seasonal changes in metal levels, because trends

Fig. 1 – Metal concentrations in leaves and stems of Phragmites australis at intertidal marshes along the river Scheldt (average ± standard deviation, n = 3).

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Table 2 – Reported metal concentrations (mg kg−1 DM, range, average or average ± standard deviation) in leaves of Phragmites australis Reference

Location

Cd

Cu 6.3–7.7 7.0

2.6–220 0.1 ± 0.0

3.2–7.8

Pb

2.7–4.1 2.2

8.0–12.5 10.0

0.3 ± 0.1

0.2 ± 0.0 <1.0

Zn

Wastewater treatment plant Control

Baudo et al. (1985) Keller et al. (1998) Larsen (1983)

Lake Mezzola, Italy Danube Delta, Romania Mid-Jutland, Denmark

Lehtonen (1989)

Espoo, Finland, Lake 1 (Hauklampi) ¨ Lake 2 (Kurkijarvi)

Vymazal et al. (2007) Lesage et al. (2007a,b)

Review natural and constructed wetlandsa Wetland treating domestic wastewater

0.03b –9.7b 0.014–0.071

0.4b –15

1.5–72b 2.2–5.7

0.6b –3.6 0.29–0.67

0.09b –264b 0.39–1.1

13–277b 20–49

Our study

Tidal marshes of the Scheldt estuary

0.03–0.218

0.28–1.21

2.7–8.6

0.5–5.8

0.54–7.09

11–57

b

2.3–3.8 0.9

Ni

Abdel-Shafy et al. (1994)

a

0.9–1.0 0.5

Cr

11–40

2.0 ± 0.7 2.0 ± 0.9

<0.1 <0.1

386–477 339

33.7 25 ± 8 14 ± 2

Outliers commonly found in wetlands receiving acid mine drainage were excluded from the dataset. Analyses of shoots, i.e. leaves + stem together.

are very different between metals. Generally, though, individual leaves acquire greater concentrations of metals over their life span. Weis et al. (2003) noted that a great variation among leaves within the same plant at any given time of sampling can be observed. This suggests that it is important to analyse a large number of leaves. It might not be assumed that a few are representative of the plant as a whole (Weis and Weis, 2004). Analytical methods used to digest the reed plant material prior to analysis also can significantly affect the observed metal concentrations (Du Laing et al., 2003). Reed plant matrices are particularly difficult to digest as they contain, e.g. high amounts of lignin and cellulose (Lenssen et al., 1999). The effects depended upon the metal and sample type. All these factors warrant great caution in comparing metal concentrations between different studies. Based on the similar order of magnitude of the concentrations, we can, however, conclude that the environmental risks related to reed plant material decomposition and consumption are similar to those at most of the other sites, for which results have been reported in the literature. Concentrations of most metals were significantly higher in the leaves than in the stems. Vymazal et al. (2007) attributed

this to the metals being mainly accumulated in leaf vacuoles. Zinc concentrations, however, were highest in the stems. This has also been observed by Larsen (1983) and Lehtonen (1989). Zinc could accumulate there as it has an essential function in the biosynthesis of the plant growth hormone indolyl-3acetic acid, which is primarily active in the stems (Schierup and Larsen, 1981a). A minimal level of 40 mg Zn kg−1 DM was observed, which might reflect a minimum level needed for the plant. However, Zn concentrations up to 140 mg kg−1 DM in stems were found at some sites. All highest levels were encountered at shorter distances from the river mouth, in the brackish part of the river mouth where sediments exhibited high chloride contents and conductivities. This suggests that the high Zn uptake is related to the higher salinity at these sites. Previous studies in the Scheldt estuary also revealed an increasing metal mobility and bioavailability with increase in salinity. Soil solution metal concentrations were shown to increase. This was attributed to the formation of soluble complexes with anions from the river water (e.g. chlorides) and/or metal release from the sediments by cationic exchange (e.g. Ca2+ , Mg2+ , Na+ , etc.) (Du Laing et al., in press). In addition, metal uptake by ground-dwelling spiders was significantly higher in the more saline intertidal marshes downstream the

Table 3 – Reported metal concentrations (mg kg−1 DM, range, average or average ± standard deviation) in stems of Phragmites australis Reference

Location

Cd

Cr

Cu

0.4–0.5 0.4

1.6–3.1 0.8

3.4–5.1 3.9

0.7–110

Ni

Wastewater treatment plant Control

Baudo et al. (1985) Larsen (1983)

Lake Mezzola, Italy Mid-Jutland, Denmark

<0.03

1.1–7 2.4

Lehtonen (1989)

Espoo, Finland, Lake 1 (Hauklampi) ¨ Lake 2 (Kurkijarvi)

<0.1 <0.1

1.5 ± 0.5 1.4 ± 0.7

Vymazal et al. (2007) Lesage et al. (2007a, b)

Review natural and constructed wetlandsa Wetland treating domestic wastewater

0.17–0.3b 0.007–0.083

4.0–33b

1.1–15.6 0.91–4.7

2.5–4.1b 0.20–0.65

0.08–2.2b 0.25–0.43

18–89 11–70

Our study

Tidal marshes of the Scheldt estuary

0.014–0.072

0.29–0.72

0.5–4.8

0.2–4.1

0.21–0.99

36–137

b

Outliers commonly found in wetlands receiving acid mine drainage were excluded from the dataset. Data originating from only two studies.

1.5–2.5 6.1

Zn

Abdel-Shafy et al. (1994)

a

1.2–2.0 2.0

Pb

<1.0

184–329 257 7–34 48 130 ± 30 55 ± 25

314

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Scheldt estuary compared to sites upstream (Du Laing et al., 2002). This trend is not observed with Cd, which generally has a similar behaviour as Zn in the environment (Smies, 1983). Whereas the mobility of Cd is also expected to increase with salinity, microbial sulphide formation in the deeper sediment layers might tend to counteract this trend (Lasat, 2002; Du Laing et al., in press). Because the solubility of CdS is orders of magnitude lower than the solubility of Zn (Lindsay, 1979), Cd will precipitate to a larger extent than Zn. Plant physiological factors might provide another reason for the different behaviour of Cd compared to Zn. Cadmium is not an essential element, and plants may be able to selectively eliminate it from their tissues. Remarkably high Pb concentrations in the leaves were observed at two sites. These sites are surrounded by roadways and industrial zones with metallurgic activities, which may probably account for elevated atmospheric Pb deposition on the leaves. Moreover, a controlled flood area is being constructed nearby. During summer, the digging and transport of dry, contaminated soil during the construction works might have increased atmospheric deposition of soil particles on the leaves. Metal concentrations in the plumes ranged from 19 to 117 ␮g kg−1 DM for Cd, from 98 to 408 ␮g kg−1 DM for Cr, from 3.1 to 7.0 mg kg−1 DM for Cu, from 0.5 to 2.3 mg kg−1 DM for Ni, from 0.4 to 4.5 mg kg−1 DM for Pb and from 36 to 132 mg kg−1 DM for Zn. Relative variations were highest for Pb, again suggesting the importance of atmospheric deposition. There were no clear relationships between total metal concentrations in the sediments and in the reed stems. Aiming to identify the environmental factors that could significantly affect metal uptake by the plants, we calculated transfer coefficients Kd for each sampling site as the ratio of the metal concentration in the leaves or stems of the reed plants (mg kg−1 DM) to the total metal concentrations in the sediments (mg kg−1 DM). Transfer coefficients of Cu and Zn are generally higher than those of the other elements (Fig. 2). Moreover, Cu and Zn transfer from sediments to stems is significantly higher at some sites near the river mouth. Sediments of these sampling sites have low clay and organic matter contents, and also low total metal concentrations. As Cu and Zn are essential elements, they seem to be more actively translocated to the aboveground plant parts at these

sites. Moreover, Zn uptake might be affected by the salinity, as mentioned before. Copper also seems to be actively transported to the leaves, whereas this is not the case for Zn. This might be explained by the functions of these metals in the plant. Copper is needed for the functioning of plastocyanin in the leaves, an enzyme which is used for electron transfer in photosynthesis, and in the enzyme cytochrome oxidase which is used during respiration, whereas Zn is mainly used for the production of growth hormones in the stems. These conclusions should be interpreted with great caution as we could not discriminate between internal uptake and external sorption to the plant despite the fact that plant parts were thoroughly washed preceding the analyses. As the proportion of external surface area to the total weight is higher for leaves compared to stems, leaves are likely to be more significantly affected by external sorption due to atmospheric deposition. However, external sorption could also affect the metal contents in the stems, as they are in closer contact with the contaminated sediments. Moreover, Windham et al. (2003) reported that only 4–20% of the aboveground metals in reed plants were found in leaf tissue, especially as stem biomass production was much higher during a major part of the growing season, from June to October. We therefore further focussed on the stems when aiming to assess environmental factors affecting the metal uptake by reed plants. For Cd and Zn, Kd coefficients for transfer from exchangeable metal fractions in the sediments (assessed by leaching with NH4 OAc and presented in Du Laing et al., 2007c) to the reed stems were plotted as a function of clay contents of the sediments (Fig. 3). Metal transfer to the stems was inversely related to the clay content. A lower sorption capacity due to a lower clay content might decrease the capacity of the sediments to immobilize Cd and Zn and thus increase the plant availability (Lasat, 2002). Moreover, as Zn is an essential element, the plants might actively promote Zn release from the sediment and accumulate Zn when the Zn concentration in the sediment is rather low, which is the case if clay and organic matter contents are low (Du Laing et al., 2007c). This may in turn interact with the Cd uptake (McKenna et al., 1993; Bunluesin et al., 2007). Finally, the less clayey sediments are expected to be susceptible to more rapidly fluctuating water table levels, which leads to occasional oxidation of sulphides and an increased metal availability. Vandecasteele

Fig. 2 – Transfer coefficients, calculated as the ratio of a metal concentration in the stems (A) or leaves (B) of the reed plants (mg kg−1 DM) to the total metal concentrations in the sediments (mg kg−1 DM), as a function of the distance from the river mouth.

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Fig. 3 – Transfer coefficients for Cd and Zn, calculated as the ratio of a metal content in the stems of the reed plants (mg kg−1 DM) to the exchangeable metal contents in the sediments (mg kg−1 DM), as a function of the clay content of the sediments (black boxes = sediments contain less than 1000 mg Cl− kg−1 DM, white boxes = sediments contain more than 1000 mg Cl− kg−1 DM).

et al. (2005b) also found lower Cd concentrations in the leaves and bark of the wetland plant species Salix cinerea with increasing duration of submersion periods in the field. Under greenhouse conditions, an upland hydrological regime resulted in elevated Cd and Zn concentrations in the leaves compared to a wetland hydrological regime. Moreover, initially submerged sediments emerging only in the second half of the growing season resulted in elevated Cd and Zn foliar concentrations at that time, whereas foliar Zn concentrations were high at a sandy-textured oxic plot with low sediment metal concentrations (Vandecasteele et al., 2005a). This effect of reducing conditions on metal availability was also reported by others (Gambrell, 1994; Du Laing et al., 2007a,b). For Zn, the inverse relationship between these transfer coefficients and the clay contents is less clear, as salinity also seems to affect the uptake. Higher than expected Zn transfer coefficients were observed for the four sites with a sediment chloride content above 1000 mg kg−1 , regardless their clay content. When sites with sediment chloride contents above 1000 mg kg−1 were omitted from the dataset, the inverse relationship becomes more clear (Fig. 3, black boxes). Chloride contents below 700 mg kg−1 (or conductivities below 1 mS cm−1 ) did not have effects on Zn uptake. No clear relationships were found between Cd and Zn transfer coefficients and chloride contents or conductivities for the complete dataset or only the freshwater part of the estuary (conductivity below 1 mS cm−1 ). The Kd coefficients for transfer of Cd and Zn from exchangeable metal fractions in the sediments to the reed

stems can be predicted by the following linear regression models: log Kd,Cd = −0.147 − 1.102 × log Clay log Kd,Zn = 1.187 − 0.804 × log Clay

(R = 0.793, p < 0.001) (1) (R = 0.632, p = 0.002)

(2)

where Kd,X represents the transfer coefficient of metal X (dimensionless) calculated as the ratio between concentrations in the stem and extractable concentrations in the sediment. Clay represents the clay content in %. The intercept was not significant for Cd, whereas it was significant for Zn (p < 0.001). This reflects the role of Zn as an essential element in the growth of reed plants, which might lead to active Zn uptake and accumulation in the stems at lower exchangeable concentrations in the sediment. The prediction of Kd,Zn was significantly improved by adding chloride concentration to the model: log Kd,Zn = 1.000 + 0.239 × Cl− − 0.756 × log Clay (R = 0.881, p < 0.001)

(3)

where Cl− represents the chloride content in g kg−1 DM. The intercept, clay and chloride contents all significantly affected log Kd,Zn (p < 0.001). Standardised regression coefficients (Beta) were 0.615 for Cl− and −0.594 for log Clay. Salinity thus again seems to promote Zn uptake, whereas increasing clay contents reduced it. Replacing the chloride content by

Fig. 4 – Transfer coefficients for Cu and Cr, calculated as the ratio of a metal concentration in the stems of the reed plants (mg kg−1 DM) to the total metal concentrations in the sediments (mg kg−1 DM), as a function of the organic matter content of the sediments.

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conductivity slightly reduces the prediction capability of the model (R = 0.857). Prediction capabilities for Cd uptake could not be significantly improved by adding chloride contents to the model, which indicates that Cd uptake by reed plants is not affected by salinity. This seems to oppose results of other authors, reporting, e.g. the mobility (Du Laing et al., in press) or plant availability (McLaughlin et al., 1994) of Cd to increase with increasing salinity. Du Laing et al. (in press), however, found that the presence of sulphides suppresses this salinity effect in wetland soils. This is especially the case for Cd as this element is usually present in very low concentrations. Therefore, the presence of only a very small amount of sulphides is already sufficient to completely precipitate Cd and mask the salinity effect. Moreover, correlation coefficients could not be improved by replacing the clay contents by organic matter contents or adding organic matter contents to the model. A similar inverse relationship was observed for Cr, Cu and Ni between transfer from the total metal pool in the sediments to the stems and sediment organic matter content (Fig. 4). As Cr, Cu and Ni have a high affinity for organic matter, a lower sorption capacity due to a lower organic matter content might decrease the capacity of the sediments to immobilize them and thus increase the plant availability. Moreover, the plants might actively promote Cu, Cr and Ni release from the sediment by the excretion of root exudates (Lasat, 2002; Xu and Jaffe, 2006) and accumulate Cu, Cr and Ni when their concentration in the sediment is rather low. This is the case if clay and organic matter contents are low, as mentioned by Du Laing et al. (2007c). Finally, higher organic matter contents may be associated with a more intense microbial activity, resulting in a more reduced state of the sediments and higher sulphide amounts less deep in the sediment. This in turn might cause metals to precipitate as sulphides and thus will reduce their availability, as was already mentioned by (Du Laing et al., 2007a,b, submitted). The coefficients for transfer of Cu, Cr and Ni from the total metal pools in the sediments to the reed stems can be predicted by the following linear regression models:

covariance between clay and organic matter contents prevents to clearly identify the factor that really matters. Prediction capabilities for Cu and Cr were markedly improved by adding chloride contents to the models (R = 0.856 and 0.884, respectively):

log Kd,Cu = −0.480 − 1.085 × log OM

(R = 0.745, p < 0.001) (4)

4.

log Kd,Cr = −1.680 − 0.730 × log OM

(R = 0.856, p < 0.001)

(5)

log Kd,Ni = −1.494 − 0.823 × log OM

(R = 0.711, p < 0.001)

(6)

The concentrations of metals in reed plants of intertidal marshes in the Scheldt estuary are low. They tend to be higher in the leaves than in the stems, with the exception of Zn. Only Zn is primarily accumulated in the stems. Uptake of metals by reed plants and their transfer to the stems was dependent on the clay and organic matter contents of the marsh sediments. The higher the clay and organic matter contents, the lower the uptake and transfer to the aboveground plant parts. This might be attributed to a higher sorption capacity, which reduces the metal availability. Additionally, high clay and organic matter contents can create reduced conditions and subsequent sulphide precipitation, which reduces the metal availability at lower sampling depths. Moreover, active metal release from the sediment and increased metal uptake might be promoted by the plants at lower total metal concentrations in the sediment, which are related to lower clay and organic matter contents. Next to the effect of clay and organic matter, salinity enhanced the uptake of Cr, Ni and Zn.

where Kd,X represents the transfer coefficient of metal X (dimensionless) calculated as the ratio between concentrations in the stem and total concentrations in the sediment. OM represents the organic matter content in %. Regression coefficients are 0.745 for Cu (p < 0.001), 0.856 for Cr (p < 0.001) and 0.711 for Ni (p < 0.001). The intercepts all significantly contributed to the model (p < 0.001 for Cr and Ni and p = 0.028 for Cu). Based on its large slope compared to the intercept, Cu transfer seems to be most importantly affected by organic matter contents. Prediction capabilities slightly declined when organic matter was replaced by clay content and they did not improve when clay contents were added to the model. This seems to confirm the role of organic matter in Cu, Cr and Ni uptake by the reed plants, although the

log Kd,Cu = −0.603 + 0.222 × Cl− − 1.054 × log OM (R = 0.856, p < 0.001)

(7)

log Kd,Cr = −1.718 + 0.069 × Cl− − 0.721 × log OM (R = 0.884, p < 0.001)

(8)

where Cl− represents the chloride content in g kg−1 DM. The intercept (p = 0.004 for Cu and p < 0.001 for Cr), log OM (p < 0.001 for Cu and Cr) and chloride contents (p = 0.001 for Cu and p = 0.017 for Cr) all were significant in predicting log Kd,Cu and log Kd,Cr . Standardised regression coefficients (Beta) in the regression model of log Kd,Cu were −0.724 for log OM and 0.421 for Cl− . For log Kd,Cr , the values were −0.844 and 0.224. This might suggest a role of the salinity in the Cu and Cr uptake by reed plants. Adding chloride contents to the model does, however, not significantly improve the prediction capability for Ni. Finally, the transfer coefficient of Pb between the total metal pool in the sediments and the reed stems can be predicted by the following linear regression model: log Kd,Pb = −2.183 − 0.547 × log Clay

(R = 0.630, p = 0.001) (9)

where Kd,X represents the transfer coefficient of metal X (dimensionless) calculated as the ratio between concentrations in the stem and total concentrations in the sediment. Clay represents the clay content in %. The correlation coefficient (0.630) is much lower compared to those of the models which were constructed for Cd, Cu, Cr, Ni and Zn. Prediction capabilities could not be improved by adding extra parameters to the model.

Conclusion

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