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Biogeochemical controls on strontium fate at the sediment – water interface of two groundwater-fed wetlands with contrasting hydrologic regimes Antoine Boyer, Mélisa Hatat-Fraile, and Elodie Passeport Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.8b01876 • Publication Date (Web): 22 Jun 2018 Downloaded from http://pubs.acs.org on July 7, 2018

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Environmental Science & Technology

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Biogeochemical controls on strontium fate at the sediment – water

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interface of two groundwater-fed wetlands with contrasting

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hydrologic regimes

4 Antoine Boyer1, Mélisa Hatat-Fraile2, Elodie Passeport1,2*

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6 7 8

Department of Chemical Engineering and Applied Chemistry, University of Toronto, 200 College Street, Toronto, M5S 3E5, CANADA

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Department of Civil and Mineral Engineering, University of Toronto, 35 St George St., Toronto,

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M5S 1A4, CANADA

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* Corresponding author: Telephone: +1 416-978-5747. Email: [email protected].

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ACS Paragon Plus Environment

Environmental Science & Technology 2

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Abstract

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Radioactive strontium (Sr) is a common groundwater contaminant at many nuclear sites. Its

13

natural retention in groundwater-fed wetlands is an attractive remediation strategy. However,

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at present, the biogeochemical mechanisms controlling Sr transport at the sediment – water

15

interface are poorly understood. In this field study, Sr fate was investigated in two wetlands

16

with contrasting vegetation and hydrologic regimes. The marsh was an open-water wetland

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with constant water table and no emergent vegetation. The swamp was vegetated with

18

fluctuating water levels and a thick mat of submerged cattail litter in the water column. High-

19

resolution porewater Sr concentrations and solid-phase sediment Sr species revealed distinct

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profiles between the two wetlands. The marsh exhibited a strongly reduced environment and

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sharp concentration peaks at the sediment – water interface. In contrast, the smaller

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concentration gradients of the swamp resulted in a reduced flux of Sr to the surface water. The

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organic fraction of the sediment dominated Sr retention compared to the inorganic iron and

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manganese oxides. However, the marsh had a significant fraction of recalcitrant Sr presumably

25

due to its incorporation into sulphur and/or carbonate minerals. These results suggest that

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vegetated wetlands with fluctuating hydrologic regimes could act as efficient sinks for Sr

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pollution.


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TOC graphic

29



30

Introduction

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Strontium is a widely distributed alkaline earth metal whose mobility is controlled by

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incorporation into minerals and adsorption onto organic and inorganic compounds. The isotope

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90

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irradiation.5 Due to its high mobility, it quickly enters groundwater resources through

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infiltration of contaminated water6,7 or the leaching of wastes,1,3,8 where nuclear facilities such

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as reactors or waste management areas are present.4 Despite stricter environmental and

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security regulations, the expected increase in the world’s total nuclear electricity production by

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20509 could result in more sites being contaminated with 90Sr. At such sites, many

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groundwater-fed wetlands are present, thus becoming a pathway of 90Sr to surface water. The

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potential to use wetlands as a remediation strategy to retain 90Sr pollution deserves further

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attention. The top 10 cm of wetland sediments, called the benthic boundary, hosts various

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macro-organisms whose exposure to 90Sr amplifies the transfer of the radionuclide through the

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food chain. To protect this key environmental compartment, a better understanding of Sr fate

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at the sediment – water interface is required.

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To date, very limited research has evaluated the fate of Sr in wetlands.10–12 Kröpfelová

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et al.11 and Šíma et al.10 investigated the inlet versus outlet Sr removal efficiency of horizontal

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subsurface flow wetlands receiving municipal wastewater. Limited removal, < 28%, was

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observed for influent Sr concentrations ranging from 234 to 840 μg/L.10 In contrast, through

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laboratory column experiments, Thorpe et al. showed that both oxic and anoxic subsurface

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sand/gravel sediments had a high capacity for Sr adsorption.13 Boyer et al.12 conducted a

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detailed characterization of the Sr adsorption and desorption potentials of various wetland

Sr is a toxic radiochemical with a half-life of 28.8 years1–4 that is only produced by nuclear fuel


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substrates. The authors reported very large Koc adsorption coefficients, ranging from 549 ± 5

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L/kg for leaf litter to 5,960 ± 150 L/kg for moss. Limited desorption was observed, in particular

54

for substrates rich in proteins.

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Wetland sediments encompass a wide range of redox and microbial conditions, which are a

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result of large quantities of organic matter and frequent fluctuations of the water level.14

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Although Sr itself is not a redox-sensitive species, remaining at its Sr2+ oxidation state below pH

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13 regardless of the redox potential,15,16 Sr fate can be affected by changes in redox conditions

59

due to its complexation with redox-sensitive species. For example, Sr sulfide can form as a

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result of hydrogen sulfide production by sulfate-reducing microorganisms.17,18 However, unlike

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other metals such as copper, arsenic, and iron, Sr sulfide is soluble.10,11 Strontium sulfate, which

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is mostly insoluble, is not expected to form in wetlands since it forms only at high pH and Sr

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concentrations.19

64

Iron(III) and manganese(III/IV) oxyhydroxides are well-known sorbents for Sr.6,20,21 However,

65

their contribution to removing Sr from sediment porewater is only temporary as Sr can be

66

released under anoxic conditions upon reductive dissolution of the oxyhydroxides.22 In addition,

67

Frohne et al.23 showed that Sr may not complex with all oxyhydroxides, and other authors have

68

suggested that Sr complexes with soil organic matter may have been mistaken as complexes

69

with oxyhydroxides.20,24 Indeed, Sr can also adsorb on organic material,12,25 as well as

70

carbonates21,26–28 and clay minerals such as montmorillonite, kaolinite, and illite.29–31 While

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Boyer et al.12 highlighted the role of organic matter and specific functional groups in the strong

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retention of Sr, to date, the relative contributions of organic matter and inorganic compounds

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for Sr retention in wetland sediments is still unclear.



74

The goal of this study was to compare the speciation of Sr at the sediment – water interface of

75

two wetlands with different vegetation and hydrologic regimes. Specifically, the objective was

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to investigate the relative roles of organic and inorganic compounds for Sr fate. The main

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components of the approach were an in situ characterization of steady-state porewater

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biogeochemical signatures and solid-phase sediment chemistry. The implications of the results

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for the management of wetlands receiving Sr-polluted water are discussed.

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Materials and methods

82

Site description

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The field study was conducted at the Canadian Nuclear Laboratories (CNL) at Chalk River

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(Ontario, Canada), located about 150 km north-west of Ottawa along the southwestern bank of

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the Ottawa River. Groundwater at the site contains natural Sr. Two groundwater-fed wetlands

86

located near the north boundary, a marsh and a swamp, were sampled. The swamp sampling

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location was approximately 130 m upstream of the marsh’s, and both presumably received the

88

same input of Sr from the groundwater. The marsh was continuously saturated. Contrarily, the

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water level in the swamp changed significantly during the sampling period, resulting in the

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alternation of flooded vs. partially saturated sediments, as seen in Supporting Information (SI)

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Figure S1. The marsh was an open water wetland with no emergent vegetation but some algae,

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while the swamp was densely vegetated with cattails (Typha latifolia) (SI Figure S2). In the

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swamp, no living plants were present in the immediate vicinity of the porewater and sediment

94

sampling locations. However, the swamp surface water had a dense mat of submerged cattail

95

litter.


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Field sampling

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The wetland sediment porewater was collected using passive equilibrium samplers called

99

peepers, allowing a 1-cm vertical sampling resolution (SI Figure S3). The peepers were covered

100

by a 0.22-μm polyvinylidene fluoride membrane (Synder Filtration, Vacaville, USA), previously

101

tested to ensure its permeability to Sr (SI Section S2 and Figure S4).

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The peepers were washed with soap and water to remove any machining residues. They were

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then soaked in 5% nitric acid for 2 days and rinsed with deionized (DI) water. The membranes

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were soaked in distilled water for 2 days and DI water for one day, replacing the water every

105

day. After cleaning, the peepers were assembled while submerged in DI water. The membrane

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was held in place between the cover and the body of the peeper by multiple nylon screws.

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After assembly, the peepers were placed in a hermetic barrel of DI water where N2 was

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bubbling to keep the water oxygen content below 4%. The peepers were stored and

109

transported in the barrel until deployment 5 days after the start of the deoxygenation process.

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A total of four peepers were deployed, two in the swamp, called Swamp-1 and Swamp-2 and

111

two in the marsh, called Marsh-1 and Marsh-2 (SI Figure S2). At each location, the two peepers

112

were placed about 30 cm away from each other. The peepers were retrieved after 5 weeks in

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the field to allow for equilibrium. Any cells that were not permanently submerged or appeared

114

damaged were not sampled. Their submersion was monitored on a weekly visit of the sites. The

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approximate relative depth of the sediment surface, defined as depth zero, for each peeper at

116

the time of retrieval was determined by visual observation of residual sediment and

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discoloration on the membrane. At the time of retrieval, the porewater samples from each



118

chamber were collected using a syringe (SI Section S1). Peeper retrieval and sample collection

119

was completed in less than an hour.

120

After the peepers were retrieved, three sediment cores were sampled at each site. The cores

121

were sampled less than 20-cm away from their respective peeper locations. The cores were

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sampled and extruded using an Incremental Core Extruding Apparatus from Aquatic Research

123

Instruments (Hope, USA). The coring tube was pushed into the sediments manually, sealed and

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the core pulled out. The length of each core was limited by the maximum manual strength that

125

could be applied to the coring equipment. On average, the peepers were 63 cm deep in the

126

swamp and 61 cm deep in the marsh, but the core depths were 40, 40, and 35 cm in the

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swamp, and 60, 45, and 45 cm in the marsh. The sediment in the swamp was unconsolidated

128

and some mixing could have occurred when the core was collected. In the swamp, the depths

129

of the sediment cores were shorter than in the marsh due to the difficulty of sampling through

130

the consolidation of organic detritus in the swamp sediment. In the marsh, the two 45-cm long

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cores were shorter than the peeper length due to an impenetrable layer of fine sediment,

132

which was shallower than in the longer marsh sediment core. The cores were extruded at 5 cm

133

intervals, air dried, and stored at room temperature until further analysis.

134 135

Porewater analysis

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All porewater samples were analyzed within one week for dissolved organic carbon (DOC) and

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anions concentrations, and within 2 weeks for metals. Porewater DOC concentrations were

138

analysed following EPA procedure 415.1-415.332 with a Fusion UV/Persulfate TOC analyzer

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(Teledyne Tekmar, Mason, USA), and a limit of detection (LOD) of 0.19 mg/L. Anions were


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quantified following EPA method 300.033 with a Dionex ICS-1600 (ThermoFisher Scientific,

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Whaltam, USA) Ion Chromatograph (IC), for nitrate (NO3−), phosphate (PO43−), bromide (Br−),

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sulfate (SO42−), and chloride (Cl−); the LODs were 0.02, 0.19, 0.02, 0.02, 0.02 mg/L, respectively.

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Total iron (Fe), manganese (Mn), calcium (Ca), aluminum (Al), and Sr were analysed following

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EPA method 60034 by Inductively Coupled Plasma Optical Emission Spectroscopy (ICP/OES) with

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a 700-series ICP/OES (Agilent Technologies, Santa Clara, USA), and LODs of 0.07, 0.0008, 0.02,

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0.05, and 0.0006 mg/L, respectively.34

147 148

Sediment analysis

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All 5-cm sediment subsamples were air-dried, ground, and sieved to 2 mm. One gram of sample

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was used for elemental analysis using a LECO SC444 (LECO, Saint Joseph, USA) for total carbon

151

(TC), total inorganic carbon (TIC), total organic carbon (TOC), and total nitrogen (TN) analyses.

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The air-dried samples were combusted to CO2 for TC analysis, and were combusted after ashing

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at 475 °C to measure the TIC. The TOC was calculated from the difference between TC and

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TIC.35 The Dumas combustion method was used for TN analysis.36

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Sequential extraction procedures (SEPs) have been widely used to speciate adsorbed metals.37

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To speciate the Sr bound to the wetland sediment, the core samples were sequentially

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extracted using a method adapted from Benitez38 and described in detail in SI Section S3 and

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Table S1. First, the total metals, including total Sr, was determined following a procedure

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adapted from Moldoveanu and Papangelakis.39 Then, four sequential extraction steps were

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completed on a separate sediment sample to quantify Sr in five fractions: F1: exchangeable, F2:

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adsorbed to organic material, F3: adsorbed to Mn oxides and amorphous Fe oxides, F4:



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adsorbed to crystalline Fe oxides, and F5: residual. The residual fraction was obtained by

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subtracting all fractions to the total Sr measured. All extracts’ Sr concentrations were measured

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using the same ICP/OES method as described above for the porewater samples.

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Finally, the particle size distribution was measured on three subsamples from each location

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with a Patrica LA-950 V2 Horiba after rewetting the samples and breaking down the aggregates

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via manual crushing and ultrasonication (SI Section S4).

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Results and discussion

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1. Surface water and porewater

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1.1 Dissolved organic carbon profiles

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The different hydrologic regimes and vegetation caused distinct DOC concentration profiles for

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the swamp and the marsh (Figure 1). The swamp surface water DOC concentrations were 25

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and 26 mg/L for Swamp-1 and Swamp-2, respectively, 1 cm above the sediment surface. The

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swamp porewater DOC concentrations then decreased to 17 and 15 mg/L in Swamp-1 and

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Swamp-2, respectively, at −45 cm. Below −45 cm, a sharp increase in DOC concentrations

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initiated up to 28 mg/L in Swamp-1 and 40 mg/L in Swamp-2. Such an increase in DOC

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porewater concentration in deep horizons of sediments was also observed in other studies.14,40

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For example, Baez-Cazull et al.14 attributed the deep porewater DOC concentration increase, 30

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cm below the sediment surface, to organic acids originating from the partial microbial

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fermentation of organic matter in this zone.


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Environmental Science & Technology

Swamp-1

Swamp-2

Marsh-1

Marsh-2

b

c

d

e

f

g

h

i

j

Depth [cm]

a

182 183

Figure 1: Porewater concentration gradients for Swamp-1 and 2 (top panels), and Marsh-1 and 2 (bottom panels). The dissolved

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organic carbon (DOC), total iron, total manganese, sulphate, and total strontium concentrations in the two peepers for each

185

separate location are plotted using appropriate scales. The straight lines between the data points were drawn to add clarity to the

186

figures and do not represent a model. Zero on the vertical axis represents the approximate depth of the sediment surface.

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11



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Similarly, in the present study, the high porewater DOC concentration measured below −45 cm

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was likely a remanence of the emergent vegetation found in the swamp. In contrast, the marsh

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DOC concentration profiles exhibited a sharp DOC concentration peak near the sediment –

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water interface, with maximum porewater DOC concentrations observed at −1 cm. The DOC

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concentration peaks were 52 and 44 mg/L for Marsh-1 and Marsh-2, respectively. This increase

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was followed by a sharp decrease down to −15 cm where DOC concentrations reached 24 and

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23 mg/L for Marsh-1 and Marsh-2, respectively. In the remainder of the marsh vertical profile,

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the DOC concentrations decreased gradually to less than 20 mg/L below −50 cm.

196

197

1.2 Redox geochemistry

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No NO3− was detected in the sediment of both the swamp and the marsh. The reducing

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conditions were therefore anoxic and past the NO3− reducing state. The swamp had lower

200

porewater concentrations of total Fe, < 13 mg/L, and total Mn, < 0.7 mg/L, than the marsh with

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up to 52 and 7 mg/L of total Fe and total Mn, respectively (Figure 1). Consistent with these

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ranges, the swamp also had higher SO42− concentrations, up to 2.7 mg/L, than the marsh, < 1.2

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mg/L. These differences between the swamp and the marsh are typical of lower concentrations

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of reduced species in the porewater of environments with fluctuating water tables.41 In both

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the swamp and the marsh, the concentrations of the redox-sensitive species showed a peak

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near or below the sediment – water interface. This peak was sharper in the marsh than in the

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swamp, due to the swamp’s fluctuating water levels that translocated the elements to lower

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depths.41,42


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In Swamp-1, the downward increase in total Fe and Mn concentrations and the decrease in

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SO42− concentrations occurred in parallel from the sediment – water interface down to −20 cm,

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indicating concurrent reduction of SO42−, Fe(III), and Mn(III/IV). A similar pattern was found in

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Swamp-2 except that the increase of total Fe was delayed until −10 cm, suggesting that the

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reduction of Fe(III) occurred after that of SO42−. The reduction of SO42− concurrently or before Fe

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is typically observed in environments with unavailable Fe, significant high concentrations of

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DOC, or high concentrations of SO42−.14 This has been observed by others.14,43,44 The

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concentrations of SO42− in the swamp were not elevated, < 2.7 mg/L, but the DOC

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concentrations, ranging from 25 to 30 mg/L were high.45 It is hypothesized that the reduction of

218

SO42− concurrently with or before that of Fe in the top 20 cm of the swamp sediment could

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have been caused by the unavailability of Fe due to its complexation with available DOC.45

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Below −20 cm, the redoxamorphic profile was similar for the two peepers located in the

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swamp: SO42− concentrations increased while total Fe and total Mn concentrations decreased.

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In this −20 to −50 cm interval, SO42− concentrations averaged 1.0 ± 0.3 mg/L. This value is above

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the threshold of 0.5 mg/L defined by Chapelle et al.,46 indicating that the sediments were likely

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still in SO42− reducing conditions and not yet undergoing methanogenesis. The decrease in Fe

225

and Mn could indicate that the redox conditions in the wetland are past the reduction of the

226

oxyhydroxides, or that Fe and Mn are precipitating with sulfides forming insoluble species.47,48

227

The increase in total Fe and total Mn concentrations observed beyond −50 cm in Swamp-1 and

228

−45 cm in Swamp-2, could indicate a zone where the reduction of the oxyhydroxides is

229

occurring. Alternatively, since these increases correspond in scale and location with the

230

increase in DOC concentration for both peepers, the increase in total Fe and Mn concentrations



231

could be a result of resuspension of Fe(III) and Mn(III/IV) complexed with organic matter

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colloids smaller than 0.22 µm, the peeper membrane pore size. Similar results were obtained

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by Pullin and Cabaniss,49 who found that Fe(III) could be complexed by fulvic acids in anoxic

234

conditions.

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In the marsh, the redoxamorphic conditions were consistent between the duplicate peepers.

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Highly compressed redox gradients were observed at the sediment – water interface. Above

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the sediment – water interface, SO42−, Fe(III), and Mn(III/IV) were reduced concurrently, with

238

rapidly changing concentrations of the redox-sensitive species overlapping a narrow vertical

239

interval. This reducing zone suggests that the 5 cm in the water column consisted of a floc-like

240

ooze layer as described by Kadlec et al.50 The ooze layer consists of a movable layer of loose

241

and unconsolidated sediments. The rapid microbial decomposition of algae-derived organic

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matter in this 5-cm deep layer, confirmed by the downward increase in DOC concentration

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could have promoted the observed anoxic conditions that led to the rapid reduction of SO42−,

244

Fe(III), and Mn(III/IV). This scenario was also observed by Baez-Cazull et al.,14 who attributed it

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to a microzone of anaerobic conditions driven by dense suspended organic material. In the top

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layer of the marsh sediment, from 0 to −5 cm, the sediments appeared to oxidize, i.e. they

247

became less anoxic, as the concentrations of total Fe and total Mn dropped while those of SO42−

248

increased, but NO3− concentrations were still below detection. The redox conditions stabilized

249

at −10 cm, where SO42− reducing conditions dominated. Similarly, as observed in the swamp,

250

the high concentration of DOC from −10 to −20 cm could have reduced the availability of the

251

oxyhydroxides contributing to the reduction of SO42− before that of Fe(III) and Mn(III/IV).

252

Between −20 and −40 cm, total Fe and total Mn concentrations increased suggesting the


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reduction of Fe(III) and Mn(III/IV). In this depth interval, the SO42− stabilised at 0.60 ± 0.07, close

254

to the threshold of 0.5 mg/L defined by Chapelle et al.46

255

Altogether, (i) the constant water level of the marsh compared to the fluctuating one in the

256

swamp, (ii) the higher porewater concentrations of total Fe and total Mn in the marsh than in

257

the swamp, (iii) the porewater Fe, Mn, and SO42− concentration profiles discussed above, and

258

(iv) the higher pH of the marsh (6.5 – 6.9) compared to the swamp (5.9 – 6.4) (SI Figure S6) all

259

point to the marsh sediments being in a more reduced state than the swamp sediments, and

260

potentially in methanogenesis.

261 262

1.3 Total dissolved strontium

263

The total dissolved Sr concentrations include the soluble Sr cation Sr2+ and any other Sr

264

complexes smaller than 0.22 µm. The porewater Sr concentrations were the highest at the

265

bottom of the sediment profiles, due to Sr-containing groundwater input in the two wetlands

266

(Figure 1). At the sediment – water interface of both locations, dissolved Sr concentrations

267

were low, < 0.06 mg/L except for Swamp-2, suggesting that the wetland sediments were able to

268

attenuate Sr and protect the overlaying surface water and benthic zone (0 to −20 cm) from Sr

269

contamination. The total porewater Sr concentrations exhibited a gradual downward increase

270

in the sediment of both locations, up to maximum concentrations of 0.19 mg/L in the swamp

271

and 0.37 mg/L in the marsh at the bottom of the profiles. While they increased significantly by

272

122% and 133% in Marsh-1 and Marsh-2, respectively, over the whole 60 cm length of the

273

marsh sediment profiles, in the swamp, total Sr concentrations only increased by 47% and 49%

274

in Swamp-1 and Swamp-2, respectively. A larger Sr concentration gradient was observed along



275

the whole sediment profile of the marsh, and in particular at the sediment – water interface.

276

This suggests that a faster upward diffusive transport of Sr to the wetland surface water was

277

occurring in the marsh compared to the swamp.

278

279

2. Sediments

280

2.1 Total organic carbon and total nitrogen

281

In the marsh, the TOC and TN concentrations were lower than in the swamp, consistent with

282

the open water of the sampling location in the marsh and the presence of dense vegetation in

283

the swamp. The TOC concentrations ranged from 0.5 to 13 % in the marsh and 15 to 24 % in the

284

swamp, while those of TN were in the range of 0.2 to 1.2 % in the marsh and 0.7 to 1.7 % in the

285

swamp (SI Figure S7).

286

Within each location, the three sediment cores showed a consistent general decreasing trend in

287

TOC concentrations, in line with the general porewater DOC concentration decrease observed

288

at both locations. This results from the decomposition of sediment organic matter throughout

289

the profile and an input of fresh organic material at the top of wetland sediments from the

290

swamp cattail litter and marsh algae. However, in one of the three cores collected in the

291

swamp, core A, an increase in TOC content was observed between −17.5 cm at 16% and −37.5

292

cm at 21%, likely due to the localized presence of organic material. There were no significant

293

changes in porewater DOC concentrations in any of the two swamp peepers at the

294

corresponding depths. It is possible that the peepers did not intersect the local organic material

295

found in core A. Alternatively, slower microbial activity, as indicated by the decrease in TN in


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core A, could have limited the breakdown of the sediment TOC into more labile dissolved

297

organic molecules.

298

The molar TOC/TN ratios were 22 ± 4 and 12 ± 2 for the swamp and the marsh, respectively (SI

299

Figure S8). The different ratios are representative of the respective sources of soil nitrogen. For

300

both sediments, a portion of the nitrogen was from microbial origin, with an expected TOC/TN

301

ratio at approximately 6.51 The low TOC/TN ratio of the marsh can be explained by the low

302

TOC/TN ratio of algae, typically between 8 ± 2 52,53 and 12.54 In contrast, the higher TOC/TN

303

ratio of the swamp compared to the marsh was due to the contribution of the emergent

304

vegetation of the swamp. Indeed, typical TOC/TN ratios of emergent vegetation range from 45

305

to 50,54 with values up to 79 ± 14 for Typha species.53 The gradual downward increase in the

306

TOC:TN ratio of the swamp sediments (SI Figure S8) could be a result of a decrease in the

307

contribution of microbial-derived nitrogen resulting from lower microbial activity.

308 309

2.2 Solid-phase manganese and iron

310

The marsh sediments, on average, had higher concentrations of Fe, by 27%, and Mn, by 76%,

311

than the swamp sediments (SI Figure S7). This is consistent with the observed swamp and

312

marsh porewater Fe and Mn concentrations, which were higher in the marsh than in the

313

swamp, by a factor of 4 and 10, respectively. Characteristic of the reduced environment, the

314

total Mn concentrations decreased with depth in both the swamp and the marsh: from 0.53 ±

315

0.06 to 0.25 ± 0.04 mg/g in the marsh, and 0.27 ± 0.03 to 0.14 ± 0.02 mg/g in the swamp. Unlike

316

Mn depletion, Fe depletion did not start at the sediment – water interface. Rather, the



317

downward decrease in Fe concentrations started between −12.5 and −22.5 cm in the swamp,

318

and between −17.5 and −47.5 cm in the marsh.

319 320

2.3 Cation exchange capacity

321

In general, the CEC decreased downward in the marsh and increased downward in the swamp

322

(SI Figure S7). The marsh CEC decrease with depth was linearly correlated with the decrease in

323

TOC content with an R2 value of 0.87; whereas, the swamp CEC and TOC content did not

324

correlate. The particle size distribution analysis revealed that neither the swamp nor the marsh

325

had significant clay content (particle size diameter < 2 μm) (SI Figure S5). However, the marsh,

326

classified as a silt loam, had finer particles than the swamp, whose texture was sandy loam. At

327

both locations, the particle size distribution drifted towards finer particle sizes with increasing

328

depth. This suggests that the main contributor to CEC was the organic matter, as also found by

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Helling et al.55 It is likely that the higher CEC observed in the swamp was a result of the higher

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TOC content, between 15 and 25%, than in the marsh where the concentration in TOC was

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Biogeochemical controls on strontium fate at the ... - ACS Publications

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