Internal Loading and Redox Cycling of Sediment Iron Explain

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Internal loading and redox cycling of sediment iron explain reactive phosphorus concentrations in lowland rivers Erik Smolders, Evert Baetens, Mieke Verbeeck, Sophie Nawara, Jan Diels, Martin Verdievel, Bob Peeters, Ward De Cooman, and Stijn Baken Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b04337 • Publication Date (Web): 24 Jan 2017 Downloaded from http://pubs.acs.org on January 25, 2017

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Environmental Science & Technology

Internal loading and redox cycling of sediment iron explain reactive phosphorus concentrations in lowland rivers Erik Smolders†*, Evert Baetens†, Mieke Verbeeck†, Sophie Nawara†, Jan Diels†, Martin Verdievel$, Bob Peeters$, Ward De Cooman$ and Stijn Baken$$

†

Division Soil and Water Management, KU Leuven, Kasteelpark Arenberg 20, B-3001

Leuven, Belgium $

Flanders Environment Agency VMM, Dokter De Moorstraat 24-26, B-9300 Aalst, Belgium

$$

European Copper Institute, Avenue de Tervuren 168 b-10, B-1150 Brussels

* Corresponding author e-mail: [email protected] Tel: +32 16 321761 Fax: +32 16 321997

1

ACS Paragon Plus Environment


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1

ABSTRACT

2

The phosphate quality standards in the lowland rivers of Flanders (northern Belgium) are

3

exceeded in over 80% of the sampling sites. The factors affecting the molybdate reactive P

4

(MRP) in these waters were analyzed using the data of the last decade (>200,000

5

observations). The average MRP concentrations in summer exceed those in winter by factor 3.

6

This seasonal trend is opposite to that of the dissolved oxygen (DO) and nitrate

7

concentrations. The negative correlations between MRP and DO is marked (r=-0.89). The

8

MRP concentrations are geographically unrelated to erosion sensitive areas, to point-source P-

9

emissions or to riverbed sediment P concentration. Instead, MRP concentrations significantly

10

increase with increasing sediment P/Fe concentration ratio (p 15, further converted here to

49

P/Fe molar ratio 10 years

6


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EXPERIMENTAL

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Monitoring data

96

The study site is the Flemish region of Belgium. The hydrology of this region is described

97

elsewhere18-20. Briefly, the region (13,600 km2) is drained to the North Sea by the catchments

98

of the Scheldt river (71%), the Meuse river (12%), the Yser river (10%) and the area of

99

Brugse polders drained by canals (7%). Major sub catchments include the Lower Scheldt

100

basin, the Demer basin and the Nete basin. Flanders has a mainly flat topography, with

101

increasing elevation in the southern part, up to 156 m above sea level. The NE part of

102

Flanders is characterized by sandy soils, the W part consists of sandy, loamy sand and clay

103

soils, while the south of Flanders has loamy soils, which can be highly sensitive to erosion

104

due to combination with slopes up to 5-10%. About 75% of surface water discharge in

105

Flanders is attributed to baseflow with a minor part being surface runoff and interflow19. The

106

Nete and Demer basins (NE Flanders) are characterised by Fe(II)-rich groundwater, with

107

average 20 mg Fe (II) L-1

108

rivers in these catchments results in rapid oxidation and precipitation of ferrihydrite,

109

contributing extensively to the authigenic sediment load of these rivers.

110

The dataset of the freshwater physicochemical monitoring network of the Flanders

111

Environment Agency VMM was consulted for the sampling period January 2003-November

112

2015. It has 3534 sampling sites distributed over Flanders and includes navigable and non-

113

navigable rivers, ditches and a small fraction (0.9%) of lentic systems. Since 2003, the

114

network had been expanded with specific points almost exclusively affected by agricultural

115

activities under the Flemish Manure Action Plan (termed ‘MAP points’). The sampling

116

procedures and water analysis use a compendium21. Briefly, grab water samples are collected

14

. Seepage of this groundwater in the mainly groundwater-fed

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under random hydrological conditions ca. 30 cm under the water surface. The sampling

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periodicity varied but the number of samples per month was fairly constant over the entire

119

period, the MRP data were available for, on average, 40% of all sample points each month.

120

Water temperature, O2, pH, and electrical conductivity (EC) are measured on the spot and

121

samples collected for other analyses stored cold. Total P (TP) is measured mainly with ICP

122

(Inductively Coupled Plasma) on unfiltered but acidified (pH 1-2) samples. The compendium

123

prescribes that orthophosphate (oPO4) is determined with the molybdate blue method

124

following an ISO guideline22. It is well established that the acid colorimetric methods for P

125

include colloidal P

126

colloidal PO4. Here, we prefer the operationally defined term molybdate reactive P (MRP).

127

The detection limits (DL) for MRP varied in the database between 0.005 and 0.025 mg P L-1

128

for MRP and MRP data of samples < DL was replaced by half of the corresponding limit

129

(10% of data). For TP, this involved 15% of data.

130

Data of P, Fe and Al of the freshwater sediment database of the Flemish Environment Agency

131

were collated. The Flemish Environment Agency has a network of 300 sediment sampling

132

sites that are sampled once every four years. The sediments are collected with a 2 L Van Veen

133

sampler to about 20 cm depth. The sediment analysis is made in aqua regia extracts.

134

Data was queried by the statistical program R (Version 3.2.0; R core team, 2015). The most

135

important packages used in R were timeDate (Rmetrics Core Team et al., 2015), chron and

136

lubridate. Maps were made with the open source QGIS software (version 2.10.1-Pisa). Maps

137

with ordinary kriging of MRP and sediment Fe and P data were made with SAGA GIS 2.0.

138

Variograms were made in SAS (SAS institute, 2011) using the variogram function for the log

139

transformed data. For MRP, the best fit was with a nested spherical model whereas for Fe and

140

P it was a spherical model. All of the map layers used here were obtained from the website of

23

, hence the term oPO4 is misleading since the method also includes

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the Agency of Geographical Information Flanders24. The Flemish Hydrographical Atlas

142

(version 01/09/2015) contained map layers for the Flemish water streams and

143

(sub)catchments. The erosion sensitivity map of the Flemish water test (‘Watertoets’) was

144

also used, more specifically the version of 20/07/2006. Lastly the Flemish digital elevation

145

model (DEM; resolution of 100x100m) was downloaded as background layer.

146

Sediment-water incubation study

147

The P-mobilization from sediments to waters was studied in laboratory conditions to separate

148

out effects of temperature, DO and sediment properties. Five sediments were collected in

149

Flanders in February 2016, i.e. in mid-winter season (Table S 1). The sediments were

150

collected from small rivers in an agricultural grassland area except the sample Bierbeek which

151

was in a grassland at 250 m downstream of a wastewater treatment plant. Some surface water

152

was included in the sediment holding container and all sediment samples were stored

153

submerged in that water at 4°C for maximally 15 days pending wet sieving (10 mm); a

154

subsample was oven-dried (70°C) and analyzed for Fe, Al and P with acid oxalate extract

155

and pH. The sediment-water incubation study was performed with these five wet sediments at

156

20° C or 4° C with or without aeration, each treatment duplicated in a complete factorial

157

design (n=40). About 300 mL water saturated sediments were gently poured into 1.1 L

158

polypropylene pots (diameter 9.5 cm) followed by the addition of 700 mL (about 7 cm height)

159

CaCl2 10-3M that had been deoxygenated by 24 h bubbling with N2 gas . The pots were

160

installed in controlled temperature rooms (±1°C) with open top, aerated treatments used

161

hobby aquarium pumps injecting air at 3 cm from the water surface through 0.2 mL pipette

162

tips. After 7, 14 and 21 days, a solution sample was collected at 1 cm above the sediment

163

surface, 0.45 µm filtered and analysed for MRP26 within 32 h; a second filtered subsample

164

was acidified pending analysis for total dissolved elements by ICP-MS (Agilent 7700x). The

25

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solution pH was measured at 3 cm above the sediment surface and the DO profile was

166

measured at 0.2-cm intervals from the top down to the sediment-water interface with a O2

167

micro sensor (tip diameter 100 µm; Unisense, Denmark) mounted on a motor-driven micro

168

manipulator.

169

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RESULTS

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Monitoring data

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The Flemish surface water can be categorized as hard (1-3 mM Ca) and pH neutral (pH 7-8;

173

Table 1). About half of the sediment samples contain >1000 mg P kg-1, indicating the legacy

174

P of past emission. The environmental quality criteria of oPO4 (here: MRP) in the Flemish

175

region range 0.07-0.14 mg P L-1, varying with stream characteristics. The analysis of the 2014

176

MRP data arranged by stream characteristics and location reveal that 82% of the sampling

177

sites exceeded the corresponding limit. The average of the >200,000 MRP observations is

178

0.49 mg P L-1.

179

The MRP data sorted by month exhibit a clear seasonal trend (Figure S 1a) with summer

180

values (July-September) exceeding winter (January-March) values by factors 1.9 (median),

181

2.8 (average) and 3.6 (P90). The MRP distribution is more right-skewed in summer than in

182

winter. The seasonal MRP trend coincides with the inverse of the seasonal DO trend, whereas

183

the seasonal water temperature trend slightly precedes these by 1-2 months (Figure 1). The

184

averages of MRP were calculated for each month 2003-2015 for the entire region (n=152) and

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these averages (log transformed) correlated more strongly with corresponding DO (r=-0.89)

186

than with temperature (r=0.81; Figure S 2). The summer increases in MRP are associated with

187

increased electrical conductivity and TP but none of both increase to the same extent as MRP.

188

Trends in nitrate concentrations are opposite to those of phosphate (MRP or TP; Figure S 1c)

189

with peaks in winter months, about 6 month after MRP peaks.

190

The MRP fluctuations were further analyzed per sampling site: the averages of MRP in

191

summer months (July-Sept., all years) were divided by corresponding winter months averages

192

(Jan-March). Sampling sites for which n 40% 11



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of the mean were excluded. The summer-winter MRP ratios exceed 10 at some points and are

194

positively related to the summer MRP, but unrelated to winter MRP concentrations (Figure S

195

3), i.e. sampling sites where large concentrations are found in summer are characterized by a

196

large MRP seasonality. The long-term average MRP concentrations are smallest in NE

197

Flanders (Figure 2) where summer-winter MRP ratios are also smallest (Figure S 4). The

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summer-winter MRP ratios are geographically unrelated to erosion sensitive areas (Figure S

199

4). Largest average concentrations are found in western Flanders where the summer-winter

200

ratios are large (Figures 2 and S 4). Along the coastline in western Flanders, summer-winter

201

MRP are smaller whereas average MRP are large in contrast with the general trend. This may

202

be explained by the groundwater P, which is markedly elevated near the coastline (not

203

shown), similar as found The Netherlands27.

204

The geographical distribution of long-term MRP data has a pronounced W-E gradient (Figure

205

2) which unlikely relates to that of P emissions. The household point emissions of P (2011

206

data28 are largest in the most densely populated center of the Flemish region (Figure S 5,

207

bottom). The livestock density is largest in western Flanders, however MRP is unlikely

208

proportional to livestock density: the current (201329) animal manure production (area based

209

values) is about 1.9-fold larger in the western West Flanders province than in the north-

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eastern Province of Antwerpen whereas the long-term average MRP concentrations differ

211

factor 3.1 (0.88 versus 0.28 mg P L-1) and agricultural emissions are not dominating the total

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P emissions28. Livestock does typically not directly interact with the surface water. The

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sediment P concentrations, used as an index of historical and current P emissions, are

214

geographically also unrelated to MRP (Figure S 5, top). In contrast, the average MRP

215

concentrations are lower where sediment Fe is large (Figure 2). A statistical approach was

216

taken to underpin these spatial associations: the MRP and sediment data were combined by 12


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dividing the database in 98 subcatchments and calculating subcatchment average values. This

218

analysis reveals no relationship between MRP and sediment P (Figure 3) but a statistically

219

significant decrease of the average MRP with increasing sediment Fe, both expressed as log

220

transformed values (p= 0.014; R=0.06, Figure S 6). The MRP correlates even stronger (and

221

positively) with increasing sediment P/Fe ratio (p=0.0006 in winter and p=0.0013 in summer),

222

the slope being steeper in summer than in winter (Figure 3). The subcatchments were divided

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according to molar P/Fe ratio using the cutoff of 0.12 (molar based), a value previously

224

reported as threshold below which the P mobilization was less likely6. The seasonality of

225

MRP is lower in subcatchments with low sediment P/Fe, whereas the reverse is true in

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subcatchments with high sediment P/Fe (Figure 4). Finally, summer MRP concentrations are

227

larger in small waterways than in larger (navigable) waterways whereas corresponding winter

228

values do not show that trend (Figure S 3). The seasonality effects on MRP concentration are

229

found in both the MAP sampling sites (mainly agriculturally affected) as the non-MAP

230

sampling sites (Figure S 3).

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Sediment-water incubation study

232

Two sediments with high Fe, but low P; were collected from NE Belgium where the MRP

233

concentrations are low; two other sediments with higher P/Fe were collected from western

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Flanders where the MRP concentrations and ratios are large; the fifth sediment was collected

235

as an intermediate between both contrasting regions (Table S 1 and Figure 2 for location).

236

The surface forced bubbling with air (aeration) did not visually induce stirring but was

237

sufficient to ensure saturated DO down to the sediment surface (>8 mg O2 L-1, with one

238

exception; Table S 2). The DO in the water layer above the non-aerated sediments readily

239

decreased with depth, effects being small in one sediment (Figure S 7). The MRP increased

240

from below detection limit in the original synthetic contact solution to values > 1 mg P L-1 in 13



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the non-aerated treatments of the two sediments with highest P/Fe ratio. No such P release

242

was found for the low P/Fe sediments or for any aerated treatment (Figure 3; Table S 2). A

243

small experiment (See supporting information: annex 1 and Table S3) including an additional

244

third treatment with N2 bubbling, showed also much larger MRP in the N2 gas bubbled

245

samples than in air bubbled treatments and that the effects of the aeration on reducing MRP

246

shown in Figure 3 are not related to dilution induced by mixing via gas bubbles. In the main

247

experiment, MRP was negatively related to DO (Figure 5) but, again, only in sediments with

248

high (>0.12) molar P/Fe ratio. The sediment P concentrations were unrelated to the MRP

249

(Figure 3). The total dissolved P (TDP) release was parallel to the release of Fe for the high

250

P/Fe sediments with a mean molar ratio (0.3) close to that of the sediment P/Fe (~0.4; Figure

251

5). In sharp contrast, the pronounced mobilisation of up to 1 mM Fe in sediment Retie (high

252

Fe but P/Fe= 0.04) did not coincide with such a release of TDP. The reddish colour noted at

253

the sediment:water interface was most pronounced in that high Fe sediment Retie suggesting

254

that any mobilized P was again sequestered at the sediment-water interface by oxidizing Fe,

255

which was present in large excess. Increasing temperature in the non-aerated treatment

256

decreased MRP and TDP release from the high P/Fe sediments, especially at prolonged

257

incubation (e.g. 21 days). The TDP concentrations were slightly above MRP concentrations

258

(median: factor 1.05) but there were a few cases where TDP well exceeded MRP. In contrast

259

to samples for TDP, those for MRP were not acidified prior to analysis to mimic the standard

260

protocol in the field monitoring. The Fe2+ oxidises quickly in pH neutral water, thereby

261

forming a precipitate at the bottom of the samples vial sequestering some P. In aerated

262

treatments, effects of temperature on TDP and MRP were small and inconsistent. The highly

263

significant temperature effect on MRP in all data collated (Table S 2) is a negative one, i.e.

264

increasing temperature decreased MRP, largely influenced by the non-aerated high P/Fe 14


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sediment treatments. The solution pH values were, on average, 0.6 pH units lower in the non-

266

aerated than in aerated treatments, likely due to the accelerated degassing of carbonic acid by

267

the forced aeration.

268

DISCUSSION

269

The MRP concentration in the lowland rivers of Flanders are probably among the largest in

270

Europe. With a general average of 0.5 mg MRP L-1 it exceeds corresponding values in other

271

EU countries by over factor 5, excluding U.K and The Netherlands for which, older (2004)

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and limited, MRP mean values are estimated 0.3-0.1 mg P L-1 respectively 15. The European

273

Environmental Agency has no recent comparative collation of MRP data but older data

274

suggest that Polish freshwaters may also contain similar MRP concentration as those in the

275

Flemish region.

276

Several lines of evidence appear from this study to propose that the temporal and spatial

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variability of MRP in the water layer in lowland rivers of Flanders is mainly related to

278

internal loading, i.e. to the legacy P in the sediment and not to the corresponding variability in

279

emission and dilution. First, the seasonality of MRP is almost reverse to that of nitrate, the

280

latter is an index of highly mobile fertilizer emissions suggesting that non-point agricultural

281

emissions are not affecting the seasonality of MRP. Second, the seasonality is unlikely a mere

282

dilution effect since conductivity responds clearly less to seasonality than MRP. Third, in

283

contrast with the common paradigm in soil science

284

sources of MRP for the entire region since erosion-sensitive soils are mainly limited to the

285

loamy belt of Flanders where MRP and its seasonal changes are average to even low (Figures

286

2 and S 4). Fourth and final, the relationships between MRP, DO and sediment P/Fe, both in

287

the field as in the laboratory incubations point to anoxic processes releasing P from the

30

erosion processes are unlikely major

15



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sediment. This occurred even in pH neutral sediments and hard waters. The anoxia as a

289

causality rather than temperature is supported by the delayed response of MRP to water

290

temperature in the field (Figure 1), the stronger association between MRP and DO than

291

between MRP and temperature (Figure S 2) and the even negative effect of temperature on

292

MRP in the incubation experiment (Table S 2).

293

Internal loading explains the spatial and temporal trends in MRP in this monitoring network.

294

This does not mean that internal loading is currently the major net source of P in this network.

295

A nutrient budget study of the Scheldt basin between 1950-2000 suggests that the sediments

296

in Flanders have clearly acted as a net sink (net retention) for P during that period17. Current

297

emissions of P have reduced from historical ones. The Flemish Environment Agency

298

estimated the point and diffuse annual P emissions in 2011 as 2600 ton P, with about 1450 ton

299

P due to point emissions 28. The annual total water discharge towards the North Sea is 4×109

300

m3 year-1

301

m3)-1 or 0.65 mg P L-1. This value is of similar magnitude as the 2010-2011 averages of MRP

302

(0.47 mg P L-1) and TP (0.77 mg P L-1). Despite uncertainties in the accuracy of diffuse

303

emissions, such budget estimate suggests a near steady state in the average sediment-water

304

exchanges of P, however with pronounced spatial and seasonal variability in the role of

305

sediments as either source or sink in the entire network. Clearly, current P emissions are

306

considerable in Flanders, and historical emissions were even larger. But redox processes

307

likely determine where and when this P accumulates in the sediment or where/when a net

308

release from it occurs.

19

. A mere dilution of the emissions in that discharge predicts 2600 ton P (4×109

309 310

The lack of elevated MRP in NE Flanders despite high sediment and groundwater P repeats

311

our earlier finding that this particular catchments (Nete and Demer) have low P/Fe ratios in 16


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the groundwater (0.03 mol P (mol Fe)-1) and stream bed through which most of the MRP is

313

fixed to ferric iron in all seasons13,

314

sediments6 qualitatively appears valid to discriminate the high MRP from low MRP releasing

315

river sediments in the laboratory and in the field. However, considerable variation remains in

316

the field data (Figure 3). Some of that variability might be related to variability in Fe

317

mineralogy in the sediments. The aqua regia extract used in the sediment monitoring network

318

is non selective, it extracts also Fe(II) sulphides and non-reducible, crystalline Fe oxides that

319

are not involved in P mobilization. Potentially a more Fe(III) selective extract, extracting both

320

the reducible Fe and the minerals that sequester P might be a more robust indicator of the

321

release of MRP. One candidate may be

322

characterizing the sediments of the incubation experiment and for characterizing P release

323

from drainage ditches13,. The release of Fe and P with a molar P/Fe ratio of 0.3 strikingly

324

corresponds with the P/Fe ratio in the oxalate extract of the corresponding two sediments, i.e.

325

0.4 (Figure 5). This suggests that the oxalate extractable Fe is the main phase to which P is

326

sorbed for these two sediments. Stoichiometric calculation show that oxalate extractable Fe

327

represents about 80% or more of the Fe+Al sorption capacity and that respiration may be

328

sufficient to completely deplete the oxalate extractable Fe, thereby releasing MRP

329

(Supporting Information, Annex 2). That calculation also suggest that the lack of P release

330

from the high Fe sediments Retie and Vorselaar is likely related to the excess ferric Fe that

331

cannot be reduced to sufficiently deplete ferric iron holding P. The two sediments with high

332

P/Fe ratio were collected in a pH neutral area and the air-dried sediments are indeed pH

333

neutral. It is commonly stated that P mobility in pH neutral, well fertilized soils and sediments

334

are controlled by Ca-phosphates, however we recently showed that Fe and Al oxyhydroxides

335

explain P mobility in such soils as good as in the more acid sandy soils (re-analysis of recent

14

. The P/Fe threshold derived previously for lake

the acid oxalate extract that was used for

17



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336

soil leaching data 31, details not shown), supporting our hypothesis that the oxalate extractable

337

Fe is the main phase to which P is sorbed for these two sediments and that Fe reduction is the

338

source of MRP release under anoxic condition. The effects of increasing temperature on MRP

339

is a negative one and might be explained by the surface oxidation of ferrous iron with traces

340

of oxygen which is faster at high than at low temperature, thereby depleting MRP and Fe at

341

prolonged incubation. Indeed, dynamics of MRP and Fe in the non-aerated treatments at the

342

20°C is characterized by a release followed by depletion of Fe and MRP, in contrast with that

343

at 4°C where the release is more constant (Table S 2).

344

The monitoring network used here is not designed to identify discharge effects on the water

345

quality parameters. On subsets of the data, we related MRP in relation to rain events and

346

found that peaks in MRP occurred after rain events in summer in an area with flat topography

347

(details not shown). Since erosion and run-off is unlikely in that area and season, such peaks

348

may be related to combined sewer overflow events or, related to that, to anoxia following

349

such discharges. Sewer overflows are active in 1.8% of their time in Flanders (2015 data).

350

These peaks logically affect the upper percentile and summer-winter ratios and require further

351

study in relation to DO trends.

352

Zwolsman

353

“anoxic waters in the low salinity zone during spring and summer makes the behavior of P

354

more complex” and predicted that the measures to restore DO will render summer dissolved P

355

“to resemble winter profiles…but the natural P buffering with retard the improvement in

356

water quality”. Taken together, the weight of evidence of our study is in line with that

357

interpretation. Internal loading likely dominates the seasonal and spatial variation in MRP

358

concentrations in surface waters. As estimated above, the sediment-water exchange of

359

phosphorus is close to a steady state for the annual average of the entire region, hence

12

concluded after several cruises on the Scheldt Estuary almost 30 years ago that

18


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sustained efforts will be needed to further reduce emissions. Our results indicate that this is

361

best accompanied by measures to avoid episodes of low DO levels in summer that lead to

362

peak MRP concentrations due to redox cycling of Fe., e.g. by further reducing the biological

363

oxygen demand in effluents. Managing MRP concentrations in water may be more

364

importantly affected by redox cycling of Fe, and hence by dissolved oxygen, than by current

365

emissions.

366 367

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ASSOCIATED CONTENT

369

Supporting Information

370

Supporting information contains additional graphs, maps and tables of the monitoring data

371

and the experiments with associated details of the methodologies where required. This

372

material is available free of charge via the Internet at http://pubs.acs.org.

373 374

AUTHOR INFORMATION

375

Corresponding Author

376

*

377

Author Contributions

378

This paper was written through contributions of all authors. all authors have given approval to

379

the final version of this manuscript.

380

The authors declare no competing financial interest.

E-mail: [email protected] (Erik Smolders)

381 382 383

ACKNOWLEDGEMENT

384

Thanks to the P-team of KU Leuven: Jessica, Ruben, Kris, Charlotte, Claudia and Daniella for

385

discussions. Bart Kerré, Kristin Coorevits and Karla Moors are thanked for practical help.

386

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Table 1. Selected properties of the Flemish freshwater and sediment monitoring network 2003-2015. n sampling P10

P50

P90

total sites

surface water MRP† (mg P L-1)

0.035

0.20

1.10

216,513

3,534

TP†

0.14

0.44

1.60

158,477

2,950

pH

6.9

7.6

8.1

23,419

3,080

Ca (mg L-1)

36

97

150

1,283

582

T (°C)

5

12

19

21,571

3,042

DO† (mg O2 L-1)

3.6

7.8

11.3

18,195

2,954

P (mg kg-1 ds)

339

990

2960

1842

762

P/Fe (molar)

0.06

0.13

0.25

668

543

sediment$

†

MRP=molybdate reactive P, TP= total P, DO=dissolved oxygen.;$aqua regia soluble

elements



Page 26 of 31

Figure 1. The seasonal patterns of molybdate reactive P (MRP, blue), dissolved oxygen (DO, green) and water temperature (red) in Flemish surface water 2003-2015, all based on averages. Note the inverted axes for DO. The standard error of the mean (not shown) is about 0.1 °C for temperature, 0.1° mg L-1 for DO and 0.01 mg L-1 for MRP.


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Figure 2. Top: long-term average MRP concentrations in surface waters over the period 2003-2015; bottom: sediment Fe concentrations. Note the log scale for both parameters. The 5



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sampling points of the river sediment for the experiments are indicated as yellow stars (top). Maps contain public information obtained under the Free Open Data License Flanders v 1.0.


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Figure 3. Top: the MRP concentrations in freshwaters are unrelated to sediment P but significantly increase with sediment molar P/Fe ratio (r=0.34, winter and r=0.32, summer). Each point is the summer or winter months long-term average 2003-2015 within 98 sub-catchments, lines are linear regressions. Bottom: P mobilization from sediment in non-aerated waters at high sediment P/Fe ratio but unrelated to sediment P, temperature effects are small (see supporting information). Sediment data in top panel are based on aqua regia extraction, bottom on acid ammonium oxalate extractable elements. 29



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Figure 4. The seasonal fluctuations in MRP for sub catchment, classified according to sediment molar P/Fe ratio, and plotted against month (1=January) and dissolved oxygen (DO). Smaller MRP fluctuations are observed in the catchments with low P/Fe. Each point is a long-term average between 2003-2015 per sampling site.

30


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Figure 5. Release of P (MRP or total dissolved P) in the sediment-water incubation study as affected by the dissolved oxygen (DO) concentration at the sediment:water layer (left) and associated to the release of dissolved Fe. Data refer to all treatments and all incubation periods (0-21 days) after sediment contact. The molar P/Fe ratio of the dissolved fractions is about 0.3 (red dashed line right) for sediments with corresponding molar ratio of 0.4 (red dots) whereas it is generally below 0.01 for sediments with molar P/Fe

Internal Loading and Redox Cycling of Sediment Iron Explain

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