Decreased DOC concentrations in soil water in forested areas in southern Sweden during 1987–2008

Science of the Total Environment 409 (2011) 1916–1926

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Science of the Total Environment j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / s c i t o t e n v

Decreased DOC concentrations in soil water in forested areas in southern Sweden during 1987–2008 Stefan Löfgren a,⁎, Therese Zetterberg b a b

Swedish University of Agricultural Sciences, SLU, Department of Aquatic Sciences and Assessment, P.O. Box 7050, SE-75007 Uppsala, Sweden IVL Swedish Environmental Research Institute Ltd. P.O. Box 5302, SE-400 14 Gothenburg, Sweden

a r t i c l e

i n f o

Article history: Received 18 November 2010 Received in revised form 4 February 2011 Accepted 9 February 2011 Available online 5 March 2011 Keywords: DOC trends Forest soils Soil water Recovery from acidification Ionic strength Swedish Throughfall Monitoring Network

a b s t r a c t During the last two decades, there is a common trend of increasing concentrations of dissolved organic carbon (DOC) in streams and lakes in Europe, Canada and the US. Different processes have been proposed to explain this trend and recently a unifying hypothesis was presented, concluding that declining sulphur deposition and recovery from acidification, is the single most important factor for the long-term DOC concentration trends in surface waters. If this recovery hypothesis is correct, the soil water DOC concentrations should increase as well. However, long-term soil water data from Sweden and Norway indicate that there are either decreasing or indifferent DOC concentrations, while positive DOC trends have been found in the Czech Republic. Based on the soil water data from two Swedish integrated monitoring sites and geochemical modelling, it has been shown that depending on changes in pH, ionic strength and soil Al pools, the DOC solubility might be positive, negative or indifferent. In this study, we test the acidification recovery hypothesis on long-term soil water data (25 and 50 cm soil depth) from 68 forest covered sites in southern Sweden, showing clear signs of recovery from acidification. The main aim was to identify potential drivers for the DOC solubility in soil solution by comparing trends in DOC concentrations with observed changes in pH, ionic strength and concentrations of Aln+. As in earlier Swedish and Norwegian studies, the DOC concentrations in soil water decreased or showed no trend. The generally small increases in pH (median b 0.3 pH units) during the investigation period seem to be counterbalanced by the reduced ionic strength and diminished Al concentrations, increasing the organic matter coagulation. Hence, opposite to the conclusion for surface waters, the solubility of organic matter seems to decrease in uphill soils, as a result of the acidification recovery. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Different processes have been proposed to explain the general trend of increasing concentrations of dissolved organic carbon (DOC) in surface waters in Europe, Canada and the US during the last two decades (e.g. Erlandsson et al., 2008; Evans et al., 2005; Monteith et al., 2007; Skjelkvåle et al., 2005). Factors such as changed hydrometeorological conditions (Erlandsson et al., 2008; Haaland et al., 2010; Sarkkola et al., 2009), reduced sulphur deposition and thereby recovery from acidification (Dawson et al., 2009; Haaland et al., 2010; Monteith et al., 2007), land cover (Laudon et al., 2009; Sarkkola et al., 2009), land use and forest management (Laudon et al., 2009; Löfgren et al., 2009b; Yallop and Clutterbuck, 2009) have been shown to affect the DOC concentrations at different spatial and temporal

⁎ Corresponding author at: Swedish University of Agricultural Sciences, SLU, Department of Aquatic Sciences and Assessment, P.O. Box 7050, S-750 07 Uppsala, Sweden. Tel.: + 46 70 69 55 177; fax: + 46 18 67 31 56. E-mail addresses: [email protected] (S. Löfgren), [email protected] (T. Zetterberg). 0048-9697/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.scitotenv.2011.02.017

scales. In a recently published article, Clark et al. (2010) assessed the proposed drivers for the DOC trends and concluded that declining sulphur deposition and recovery from acidification, increasing the organic matter solubility, is the single most important factor for the long-term DOC concentration trends in surface waters. They also remarked that the variability between sites is influenced by a multitude of spatial and temporal factors. In forested and peat covered catchments, the DOC in surface waters is primarily of allochtonous origin (Thienemann, 1925). Hence, the observed DOC trends are the results of processes along the hydraulic flow paths in the soil and surface water system. For example, the solubility of soil organic matter could have increased in the soils or a larger fraction of the soluble DOC fraction could have been transported to streams and lakes. If the acidification recovery hypothesis is correct, the DOC solubility should increase both in the soil solution as well as in the surrounding surface waters as soil pH rises and ionic strength declines due to the reduced input of SO2− 4 ions. However, in contrast to the surface water trends, the soil water DOC concentrations have developed in divergent directions at a few European research sites with data from the middle of the 1990s.

S. Löfgren, T. Zetterberg / Science of the Total Environment 409 (2011) 1916–1926

Löfgren et al. (2010) found that 16, 9 and 2 lysimeters at two integrated monitoring (IM) sites in southern Sweden exhibited no change, decreasing and increasing DOC concentrations, respectively. The soil solution, which was collected along hill slopes at soil depths of 13–50 cm and in various soil types (podzols, gleysols and histosols), indicated no change or increased coagulation of DOC in the upper soil horizon during the period 1996–2007. Similar results were obtained in Norway, with no change or decreasing DOC trends during the period 1996–2006 in soil water (nsites = 18) at 15 and 40 cm soil depth in podzols (Wu et al., 2010). However, results from two sites in the Czech Republic indicate increased DOC solubility and enhanced soil water concentrations during the period 1994–2007. The soil water was collected under the forest floor at Lysina and in the mineral topsoil at Pluhuv (Hruska et al., 2009). In the UK, short-term (2000– 2005) soil water studies also show positive DOC trends at 5–20 cm soil depth (n = 9, moorlands and forests, Buckingham et al., 2008). Wu et al. (2010) suggested that small changes in the atmospheric deposition during the investigation period could explain the diverging DOC trends in soil and surface waters. However, they also suggested competition between mineral anions and DOC for adsorption sites on oxide surfaces, as a cause for the simultaneous decrease of the DOC and SO2− concentrations. The UK soil water time series were 4 considered too short for making a coupling to the surface water DOC trends (Buckingham et al., 2008). Löfgren et al. (2010) used a geochemical model to evaluate the chemistry behind the observed DOC trends in soil water recovering from acidification. According to their results, the decrease in ionic strength counteracted the increased net charge of soil organic matter (SOM) following the pH increase. This lowered the SOM net charge and thereby the DOC solubility. They concluded that the recovery from acidification does not necessarily have to generate increasing DOC concentrations in soil solution. Depending on changes in pH, ionic strength and soil Al pools, the trends might be positive, negative or indifferent. Based on time series comparisons, Hruska et al. (2009) also concluded that reduced ionic strength, rather than acidity, is the most important driver for the DOC changes. However, in their studies the decreased ionic strength covaried with increased DOC concentrations in both soils and surface waters. Hence, the ionic strength is proposed to affect DOC solubility in different directions in the studies by Löfgren et al. (2010) and Hruska et al. (2009). One explanation might be that the soils of Central Europe have been subjected to higher loads of acid deposition compared with southern Sweden resulting in lower pH and higher ionic strength as well as higher concentrations of Aln+ and SO2− in the soils. At pH 4 values below ca 4, the aluminiumoxyhydroxides are dissolved, probably increasing the DOC solubility (see Löfgren et al., 2010 and references therein). Another explanation might be that the soil water transects at the two integrated monitoring sites are non-representative for southern Sweden, which historically is the most polluted part of the country. In this region, soils and surface waters show clear signs of large-scale recovery from acidification since the early 1990s (Karltun et al., 2003; Löfgren et al., 2009a; Skjelkvåle et al., 2005). In this study, we test the acidification recovery hypothesis by Clark et al. (2010) on soil water data from 68 forest covered sites in southern Sweden. The main aim is to identify potential drivers for the DOC solubility in soil solution by comparing trends in DOC concentrations with observed changes in pH, ionic strength and − − concentrations of Aln+ as well as major anions (SO2− 4 , Cl and NO3 ). A second aim is to analyse whether the soil water DOC trends found by Löfgren et al. (2010) are representative for larger areas in southern Sweden. 2. Methods Within this project, soil water data from the Swedish Throughfall Monitoring Network (60 sites, www.krondroppsnatet.ivl.se) (Hallgren-

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Larsson et al., 1995) and reference plots to the long-term forest soil liming (4 sites, Zetterberg et al., 2006b) and ash application (4 sites, Zetterberg et al., 2006a) experiments were used. Data from tree covered sites located south of latitude 61° were selected (Fig. 1). Norway spruce (Pinus sylvestris) or Scots pine (Picea abies) is the dominant tree species with glacial till as the prevailing parent material at all sites. Between the sites, annual mean temperature and precipitation (1961–90) varied in the ranges 2.5–7.5 °C and 550–1050 mm (Swedish Meteorological and Hydrological Institute), respectively, while the total sulphur and nitrogen depositions during 2006–2008 (MATCH-model data, Swedish Meteorological and Hydrological Institute, SMHI) varied in the ranges 2–7 and 4–13 kg ha−1 year−1, respectively (Table 1). In the late 1980s, the sulphur deposition was more than twice these levels and the nitrogen deposition approximately 1 kg ha−1 year−1 higher (Westling and Lövblad, 2000). Soil water data cover the period 1987–2008 and only sites with more than 10 years of data were used. In order to avoid installation effects, data from the first year after installation were removed. Based on methodological studies within the region and with similar ceramic cups (see below), it has earlier been shown that such effects are faded away after the first winter after installation and that no ageing effects could be found after 12 years (Fölster et al., 2003). Within this project, the first lysimeter installations were done in 1986. In addition, data later than January 2005 were removed for 4 stations due to potential effects of excessive storm-felling. The methodology differs slightly between the three studies from which data has been collected. If available, soil water was collected with suction lysimeters (ceramic cups, P80, 1 μm cut-off) at 25 cm (n = 4) and 50 cm (n = 64) soil depth in recharge areas 3 times per year; before, during and after the vegetation season. At each site, 2 (liming and ash experiments) and 5 (Swedish Throughfall Monitoring Network) lysimeters are installed, but due to temporally and spatially heterogenic moist conditions, 1–5 lysimeters collected soil solution at each sampling occasion. Underpressure (−0,3 bar corresponding to approximately 550 mmHg) was created with a hand pump and the samples were collected two days later. During that time and at 23 sites, the collecting plastic bottles were stored in darkness and at ambient air temperature on top of the soil, protected by a nontransparent plastic tube (Table 3). At the rest of the sites, the samples were collected in plastic chambers inserted in the topsoil (down to 20–25 cm soil depth). The sub-samples from each site were combined into a composite sample in the field. The Swedish Environmental Research Institute, IVL and subcontractors analysed the samples using Swedish standard methods (www.slu.se/vatten-miljo/vattenanalyser). pH was analysed with combination of pH-electrodes adapted for low ionic strength waters (Radiometer) and alkalinity was titrated to pH = 5.4, which in the pHrange relevant for this study mainly corresponds to the bicarbonate (HCO− 3 ) concentration when the alkalinity is above zero. Samples for total organic carbon (TOC) analysis, hereafter termed dissolved organic carbon (DOC) due to the lysimeter filtration (0.8 μm cutoff), were measured at the Department of Aquatic Sciences and Assessment, SLU, using a Shimatzu TOC 5050 analyzer with sample injector following acidification. Major cations were analysed by ion cromathography (Dionex ICS-1100) and strong acid anions by ion cromathography (Dionex ICS-2000). Total aluminium (Alt) and, if enough sample volume was available (73% of 2385 samples), noncationic Al (Alo, Driscoll, 1984) were analysed with AAS until 2004 and thereafter by ICP-AES, while cationic Al (Ali) was calculated as the difference between Alt and Alo (Ali = Alt − Alo). In the liming and ash experiments, only Alt was analysed. Acid neutralizing capacity (ANC) was calculated according to Reuss and Johnson (1986), while ionic strength was calculated from − n+ 2− − − the H+, Ca2+, Mg2+, Na+, K+, NH+ 4 , SO4 , NO3 , Cl , HCO3 and Al concentrations. The average, net charge for aluminium (n+Al) was estimated from the relation between pH and the Ali/Alt ratio based on the median values for each site and with the assumption that Ali

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S. Löfgren, T. Zetterberg / Science of the Total Environment 409 (2011) 1916–1926

Fig. 1. Locations of the soil water sampling sites (n = 68) in southern Sweden.

corresponds to Al3+ (n+Al = Ali/Alt ⁎ 3 = (2464–0366 ⁎ pH) ⁎ 3, n = 60, r2 = 0,54, p b 0.001). By this simple method it was possible to roughly estimate the n+Al and ionic strength for occasions when Al speciation data were missing, including the eight liming and ash experiment sites. Due to the pH differences between sites, the n+Al range is 0.1– 2.8 (median 2.1, 25-percentile 1.6, n = 68). The non-parametric Seasonal Kendall test (Hirsch and Slack, 1984; Loftis et al., 1991) was used for detecting monotonous trends in the chemical time series. It was visually determined whether the trends were monotonous or not. Thiels slope (Helsel and Hirsch, 1992) was used to quantitatively estimate the trends. The trends were defined as statistically significant if the probability was less than 0.10 (p b 0.10). 3. Results The Swedish Throughfall Monitoring Network, as well as the liming and ash application experiments, was set up on sites mainly covered with semi-mature to mature (48–125 years) stands of Norway spruce and Scots pine covering large production spans with site indexes (SIH100, estimated tree height at 100 years) in the range G22–G34 (Norway spruce) and T22–T28 (Scots, pine, Table 1). Since focus is on throughfall measurements, tree-covered conditions are a prerequisite. The studies are performed on private land, however, and the owners decide upon clear-felling and other forestry operations. In 2005 and 2007, two massive storms hit southern Sweden, severely damaging 4 of the sites. This, in combination with different financing, cause the number of study sites to change with time. Table 2 shows the lysimeter installation period, the last year with data, the time

series length used for the trend analyses (the first year of data is omitted due to potential installation effects) and the number of sites in each class. Approximately 75% of the sites (n = 52) have 10–15 years data and the lysimeters were mainly installed during the period 1991–97 (Table 2). In this category, the last year of data is from the period 2005–08 for most sites. The rest of the sites (n = 16) represent data from 16 to 22 years and the time series mainly start in the end of the 1980s. The mean study period is 13.2 and 13.6 years at Scots pine (n = 13) and Norway spruce (n = 47) sites, respectively. Depending on variable and site, the number of observations varies between 12 and 62 (medians 29 and 31). The low numbers of observations are from the lime and ash experimental sites (Table 3). pH and ionic strength, which is calculated from 11 different variables, exhibit the largest and lowest numbers of observations, respectively. The soil water chemistry at 50 cm soil depth is widely varying from almost neutral conditions to very acidic sites (Table 3). Median pH and Alt varied in the range 4.18–6.67 and 0.14–5.11 mg Alt l−1 respectively, reflecting a large variation in median ANC (− 472 to 265 μeq ANC l−1). Ca2+ (nsite = 31) and Na+ (nsite = 29) were the dominant base cations and the median concentration ranges were large, 18–378 μeq Ca2+ l−1 and 62–858 μeq Na+ l−1, respectively. The Na+ dominated sites are found on the western side of southern Sweden, where sea salt is an important ionic source. The marine influence was also reflected in higher median Cl− and Mg2+ concentrations. Sulphate was, however the dominant mineral acid − anion at most sites (nsite = 41). Both SO2− 4 and Cl showed very large −1 median concentration intervals 49–708 μeq SO2− and 23–681 μeq 4 l

S. Löfgren, T. Zetterberg / Science of the Total Environment 409 (2011) 1916–1926

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Table 1 Site. latitude, longitude, altitude (m above sea level), mean temperature 1961–90 (°C, SMHI data), mean precipitation 1961–90 (mm, SMHI data), tree species (NS = Norway spruce, SP = Scots pine), planted year, site index (SIH100, estimated tree height in m at 100 years, G = Norway spruce, T = Scots pine), total sulphur (S-dep) and nitrogen (N-dep) deposition (kg ha−1 year−1, SMHI MATCH data for the period 2006–2008), and nd = no data. Site

Latitude

Longitude

Altitude (m.a.s.l.)

Mean temperature 1961–90 (°C)

Mean precipitation 1961–90 (mm)

Tree species

Planted (year)

SIH100

S-dep (kg ha−1, year−1)

N-dep (kg ha−1, year-1)

A01 A05 A21 A24 A35 A40 A44 A54 A90 A92 D11 D14 D51 D52 E02 E03 E21 E22 F09 F21 F22 F23 G09 G21 G22 H03 H21 H22 K01 K03 K10 K13 L05 L07 M13 N01 N17 O01 O05 O35 P02 P52 P70 P92 P93 P94 R09 S01 S05 S15 S16 S22 S23 T02 T03 T10 U02 U03 U04 U05 12_25* 12_50 29_25* 29_50 1000B* 2000A 2000B* 4000A

59° 59° 59° 59° 59° 59° 59° 59° 59° 59° 58° 59° 59° 58° 58° 58° 58° 58° 57° 57° 57° 57° 56° 56° 57° 56° 58° 56° 56° 56° 56° 56° 56° 56° 55° 57° 56° 58° 57° 58° 59° 57° 58° 57° 57° 58° 58° 60° 59° 59° 59° 59° 60° 59° 59° 59° 59° 59° 59° 60° 57° 57° 57° 57° 57° 57° 57° 57°

18° 18° 17° 17° 17° 18° 17° 18° 18° 17° 16° 16° 17° 17° 16° 15° 15° 15° 13° 14° 14° 15° 15° 15° 14° 16° 16° 15° 14° 14° 15° 15° 14° 13° 13° 12° 13° 11° 12° 11° 12° 12° 13° 13° 13° 11° 13° 12° 13° 13° 11° 12° 13° 14° 14° 14° 15° 16° 15° 15° 14° 14° 14° 14° 14° 14° 14° 14°

38.2 26.2 25.7 35.5 42.4 9.1 63.5 33.4 41.5 37.5 50.2 39.5 31.7 29.6 128.8 167 174.1 141.6 219 230.1 288.2 239.8 154 225.8 191.6 22.8 132.1 151.1 134.3 100.5 104.6 64 122 119.5 111 93 157 143.8 99.8 116.3 101.4 84.4 149.8 190.5 309.6 164.3 88.1 320.8 55.7 65.8 192.3 349.8 367.7 198 200.6 87.8 178.6 53.5 42.5 175.6 241.3 241.3 241.3 241.3 241.3 241.3 241.3 241.3

5.5 5.5 5.5 5.5 5.5 6.5 5.5 5.5 5.5 5.5 5.5 5.5 5.5 5.5 5.5 5.5 5.5 5.5 5.5 5.5 5.5 5.5 6.5 5.5 5.5 6.5 6.5 6.5 6.5 6.5 6.5 6.5 6.5 6.5 7.5 6.5 6.5 6.5 6.5 6.5 5.5 6.5 6.5 5.5 5.5 5.5 6.5 2.5 5.5 5.5 4.5 4.5 2.5 4.5 3.5 5.5 4.5 5.5 5.5 4.5 5.5 5.5 5.5 5.5 5.5 5.5 5.5 5.5

650 650 650 650 650 650 650 650 650 650 650 650 650 650 650 650 650 750 850 650 650 650 650 750 850 550 750 750 850 750 750 750 750 950 750 1050 1050 950 1050 950 850 1050 750 1050 950 1050 650 750 750 750 850 850 850 850 850 750 750 650 650 750 750 750 750 750 750 750 750 750

SP NS NS SP NS NS NS NS NS SP NS SP NS NS NS NS NS SP NS SP NS NS NS SP NS NS SP NS NS SP NS NS NS NS NS SP NS SP NS NS NS NS NS NS SP NS NS NS NS NS NS NS NS NS SP NS NS NS NS NS NS NS NS NS NS NS NS NS

1930 1906 1933 1902 1900 1898 1892 1885 1907 1935 1930 1933 1933 1924 1920 1897 1935 1935 1925 1950 1952 1951 1954 1920 1925 1916 1937 1937 1927 1935 1931 1931 1955 1943 1958 1923 1937 1947 1891 1919 1906 1916 1932 1930 1948 1927 1941 1930 1931 1919 1930 1951 1952 1948 1946 1924 1907 1910 1920 1927 1962 1962 1962 1962 1962 1962 1962 1962

T24 G26 G26 nd G28 G26 G24 nd nd T22 G28 T28 G30 G26 G24 G30 G32 T24 G30 T28 G28 G32 G32 T22 G28 G28 T26 G32 nd T23 G28 G34 G32 G34 G34 T24 G34 T26 G22 G26 G28 G33 G31 G30 T24 G28 G30 nd nd nd nd G32 G26 G32 T25 G28 G26 nd G28 nd G32 G32 G32 G32 G32 G32 G32 G32

3.5 3.5 2.9 3.5 3.4 3.5 4.0 3.0 3.5 3.0 3.4 2.6 2.9 3.6 2.9 2.9 2.6 2.7 3.9 2.8 2.8 3.3 4.4 4.4 3.7 4.5 3.1 4.5 4.5 4.9 5.7 4.9 4.5 4.8 6.7 5.4 5.3 4.9 4.3 4.9 4.4 5.4 3.3 3.8 2.9 4.9 3.2 2.0 4.4 4.1 4.7 3.5 2.2 2.9 3.1 2.9 2.6 2.6 2.6 2.5 3.8 3.8 3.8 3.8 3.8 3.8 3.8 3.8

7.2 7.2 6.7 7.2 7.0 7.2 7.8 6.1 7.2 6.1 7.0 6.0 6.7 6.9 6.5 6.5 6.6 6.9 9.4 6.6 6.6 7.1 9.5 9.5 8.4 9.4 6.5 9.6 10.5 10.7 10.9 10.7 10.5 11.2 13.0 11.4 12.2 9.5 9.9 9.5 8.4 11.4 8.7 9.8 7.7 9.5 8.1 4.0 8.4 7.1 8.3 6.2 4.3 6.1 5.6 6.6 5.5 6.1 6.0 5.0 8.1 8.1 8.1 8.1 8.1 8.1 8.1 8.1

34′ 3″ N 22′ 41″ N 34′ 24″ N 26′ 13″ N 5′ 45″ N 17′ 59″ N 11′ 35″ N 48′ 25″ N 35′ 12″ N 40′ 26″ N 57′ 7″ N 18′ 29″ N 17′ 58″ N 46′ 40″ N 39′ 47″ N 11′ 26″ N 9′ 11″ N 45′ 49″ N 12′ 12″ N 52′ 40″ N 50′ 7″ N 30′ 41″ N 40′ 56″ N 53′ 32″ N 3′ 39″ N 51′ 10″ N 3′ 45″ N 38′ 25″ N 23′ 21″ N 20′ 37″ N 22′ 43″ N 16′ 11″ N 10′ 33″ N 8′ 5″ N 37′ 7″ N 30′ 53″ N 22′ 6″ N 32′ 20″ N 41′ 9″ N 26′ 14″ N 1′ 46″ N 31′ 29″ N 2′ 33″ N 24′ 13″ N 46′ 28″ N 43′ 20″ N 37′ 38″ N 52′ 19″ N 0′ 32″ N 30′ 16″ N 34′ 9″ N 49′ 16″ N 35′ 47″ N 35′ 12″ N 52′ 49″ N 12′ 44″ N 55′ 51″ N 39′ 13″ N 20′ 49″ N 8′ 8″ N 8′ 44″ N 8′ 44″ N 8′ 44″ N 8′ 44″ N 8′ 44″ N 8′ 44″ N 8′ 44″ N 8′ 44″ N

3′ 33″E 7′ 23″ E 38′ 49″ E 52′ 52″ E 37′ 55″ E 24′ 52″ E 57′ 55″ E 14′ 59″ E 21′ 22″ E 58′ 13″ E 58′ 51″ E 6′ 51″ E 7′ 35″ E 13′ 21″ E 6′ 41″ E 30′ 32″ E 25′ 38″ E 8′ 42″ E 58′ 28″ E 44′ 57″ E 59′ 38″ E 20′ 34″ E 8′ 44″ E 6′ 40″ E 22′ 34″ E 19′ 16″ E 6′ 27″ E 37′ 25″ E 37′ 40″ E 58′ 48″ E 43′ 9″ E 26′ 32″ E 14′ 58″ E 30′ 27″ E 26′ 34″ E 14′ 43″ E 5′ 43″ E 43′ 27″ E 29′ 3″ E 44′ 9″ E 36′ 23″ E 28′ 44″ E 5′ 5″ E 6′ 12″ E 44′ 41″ E 58′ 23″ E 46′ 35″ E 40′ 33″ E 7′ 0″ E 29′ 8″ E 44′ 19″ E 54′ 26″ E 6′ 59″ E 42′ 28″ E 25′ 45″ E 49′ 2″ E 46′ 46″ E 15′ 6″ E 56′ 13″ E 54′ 13″ E 46′ 12″ E 46′ 12″ E 46′ 12″ E 46′ 12″ E 46′ 12″ E 46′ 12″ E 46′ 12″ E 46′ 12″ E

Cl− l−1, respectively. The NO− 3 concentrations were mostly below −1 detection limit (b0.1 μeq NO− ) and only two sites exhibited 3 l −1 significant levels (137 and 157 μeq NO− ), indicating net nitrifica3 l

tion in the soils. The large span in ion concentrations for the studied sites causes large differences in ionic strength, ranging between 0.2 and 2.5 mmole l−1 (Table 3). Among the anions, SO2− gave the 4

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S. Löfgren, T. Zetterberg / Science of the Total Environment 409 (2011) 1916–1926

Table 2 Lysimeter installation period, the last year with data, the time series length (years) used for the trend analyses and the number of sites (nsites) in each class. Lysimeter installation period

nsites

Last year with data

nsites

Time series length⁎ (years)

nsites

1986–1990 1991–1995 1996–1997

20 23 25

1996–2000 2001–2005 2005–2008

3 17 48

10–15 16–20 21–22

52 13 3

⁎ Time series length after that the first year of data is omitted.

dominant impact on the ionic strength at 65 sites, while 3 coastal sites were dominated by Cl−. For the cations, Aln+, Ca2+ and Na+ gave the most important tribute to the ionic strength at 23, 23 and 17 sites, respectively, while Mg2+ dominated at 5 sites. Thus, the study sites represent a large variation in acidity, ionic strength and ion composition especially regarding the cations. The DOC concentrations were much less variable compared with acidity and major inorganic ions, but still showed a substantial variation between the sites. Based on median concentrations, the range was 2.6–23.5 mg DOC l−1 (Table 3). At 50% of the sites, the median DOC concentration was above 8.2 mg DOC l−1. This must be considered high, taking into account that the samples are collected in podzols at 50 cm soil depth, presumably in the lower B-horizon (see Discussion). High DOC concentrations in shallow groundwater (b1 m soil depth) are earlier documented in podzols in southern Sweden, indicating insufficient DOC coagulation in the mineral soil (Löfgren and Cory, 2010). The soil solution composition was more acidic in the Norway spruce (NS) stands compared with the Scots pine (SP) stands. Statistically significant differences (p b 0.05, ANOVA) were found for pH (pHNS = 5.0, pHSP = 5.2), ionic strength (ISNS = 0.9 mmol l−1, ISSP = 0.6 mmol l−1) and Alt (Alt, NS = 1.3 mg l−1, Alt, SP = 0.7 mg l−1). A large proportion (72%) of the sites showed statistically significant (p b 0.10) decreasing SO2− concentrations with time in 4 the soil solution (Table 4, Fig. 2). The annual reductions were in the −1 order of 2–42 μeq SO2− year−1. Only one site exhibited an 4 l increasing trend. In contrast, the Cl− concentrations did not change by time at most sites (66%, Table 4). However, 10 of the sites near the coast showed increasing trends (6–48 μeq Cl− l−1 year−1), while 13 mainly inland sites exhibited decreasing Cl− concentrations (1– 36 μeq Cl− l−1 year−1, Fig. 2). Except for three sites, the NO− 3 concentrations were indifferent by time (Table 4). The reduced anion concentrations were to a large extent balanced by primarily reduced Ca2+ (data not shown) and Aln+ concentrations, but not always to such an extent that ANC remained constant by time (Fig. 2). The buffering capacity increased between 2 and 20 μeq ANC l−1 year−1 at 47% of the sites and only three sites exhibited decreasing ANC. pH increased at 22 sites due to the increased ANC, but at 54% of the sites no changes were detectable. In contrast, pH decreased at 9 sites. The pH increases corresponded to a reduction of 0.1–3.7 μeq H+ l−1 year−1 (Tables 4 and 5, Fig. 2). Over the entire study period, this corresponds to a median pH increase of 0.3 pH-units. Five sites showed a pH-increase in the range 0.5–1 pH-units (Table 5). The Alt concentrations decreased at 23 sites with 0.01–0.18 mg Alt l−1 year−1 and increased at 5 sites (Tables 4 and 5, Fig. 2). Besides affecting ANC and pH, the changed ionic composition decreased the ionic strength at 33 sites within the range 0.004–0.129 mmole l−1 year−1 (Tables 4 and 5, Fig. 2). Hence, reduced ionic strength was much more common than increased pH and reduced Alt. The main effect of changed soil solution acidity, ionic strength etc. on the dissolved organic matter was primarily no (47%) or decreased (46%) DOC concentrations (Table 4). Only 5 sites exhibited statistically significant (p b 0.10) increasing DOC trends. At 4 of these sites the median pH was in the range 5.0–5.8, while one site had pH = 4.5 (Tables 3 and 4). The annual DOC reduction rates were in the range

0.1–1.5 mg DOC l−1 year−1 (Table 5) and sites with diminished DOC concentrations were found in the entire region (Fig. 2). Fig. 3 shows the long-term changes in DOC, pH, ionic strength and Alt at the P52 site, which is an example of a site where the DOC concentrations have decreased by time. The most pronounced changes in DOC, ionic strength and Alt occurred during the period 1990–98, while pH showed a small positive trend throughout the investigation period. 4. Discussion During sampling, temperature differences might have occurred between the two types of soil water samplers. However, the temperatures in shallow soils and air are generally fairly similar in forests with closed canopies. At the Swedish IM site Aneboda (Löfgren et al., 2010) in the south-centre of our study area, the average temperature in July (1996–2009) was 16,5 °C at 2 m above soil surface and 12,6 °C at 32 cm soil depth (unpublished data). In spring and autumn the temperature differences were less and in autumn often higher in the soil than in the air. Hence, we do not believe that temperature differences between samplers have had any substantial effects on the DOC trends, since they are estimated from all three seasons. Additionally, statistically significant DOC trends were found in both sampling types (Tables 3 and 5). Scientifically, it is not appropriate to apply geochemical modelling on the soil solution data used in this study. The chemical analyses were performed on composite samples, representing a mixture of soil water chemical conditions in the lysimeters at the time the subsamples were collected. Thus, the analysed ionic composition does not mirror the actual concentrations found in nature. The sampling technique might also affect the non-parametric statistical trend analyses by giving sub-samples with significant, monotonous temporal changes in ion concentrations a greater influence compared with those maintaining highly variable or rather stable chemical conditions. Additionally, mixing of sub-samples with positive and negative trends might theoretically outbalance each other. However, studies of soil water chemistry from individual lysimeters located in the E-horizon, B-horizon and peat soil show that the soil water DOC and ion concentration trends are indifferent or in the same direction in most lysimeters (n = 27, Löfgren et al., 2010). The problem of counterbalancing should therefore be considered small or negligible taking into account the uniform soil type (podzol) and short distances (b30 m) between the sub-samples at each site. Finally, trend amplification by one or more lysimeters might affect the rate of the temporal change but not the trend direction. Therefore, the trend rates presented (Tables 4 and 5) are only indicative and will not be further discussed. Instead, we focus on the trend directions. As expected from other long-term soil water studies in Sweden (Fölster et al., 2003; Löfgren et al., 2001, 2010), Norway (Wu et al., 2010) and the Czech Republic (Hruska et al., 2009), most sites showed decreasing SO2− concentrations. These reductions are coupled to 4 reduced sulphur emissions during the last decades and an indication of soil systems in a state of recovery from acidification. This could be detected as increased ANC at almost half of the sites and increased pH and/or reduced Alt at approximately one third of the sites. These sites, showing signs of acidification recovery, are in the following used to test the DOC hypothesis by Clark et al. (2010). The surface potential of a charged colloid may be the single most important factor determining its dispersion into the water phase (e.g. Weng et al., 2002). A high surface potential increases the solubility due to more interactions with water molecules. In turn, the humic charge and soil water DOC solubility are to a large extent determined by pH, ionic strength and the soil Al pools (c.f. Löfgren et al., 2010 and references therein). Table 6 shows the number of sites with signs of recovery from acidification in terms of statistically significant (p b 0.10) positive trends in pH, negative trends in ionic strength and/or decreased aluminium concentrations. For each variable, the

S. Löfgren, T. Zetterberg / Science of the Total Environment 409 (2011) 1916–1926

1921

Table 3 Site, sample collector type, data period, min and maximum number of observations (nmin and nmax) for the water chemical variables, median concentrations of TOC, pH, ionic n+ strength (IS), total aluminium (Alt), mean charge of cationic Al (n+ ) in soil solution at 50 cm soil Al) and the major inorganic ionic constituents including cationic aluminium (Al depth at 68 sites, collected during the years 1987–2008 in southern Sweden. OS = plastic bottle on top of the soil, and IS = plastic chamber in the topsoil (20–25 cm). Site

Collector type

Data period

nmin

nmax

DOC mg l

A 01 A 05 A 21 A 24 A 35 A 40 A 44 A 54 A 90 A 92 D 11 D 14 D 51 D 52 E 02 E 03 E 21 E 22 F 09 F 21 F 22 F 23 G 09 G 21 G 22 H 03 H 21 H 22 K 01 K 03 K 10 K 13 L 05 L 07 M 13 N 01 N 17 O 01 O 05 O 35 P 02 P 52 P 70 P 92 P 93 P 94 R 09 S 01 S 05 S 15 S 16 S 22 S 23 T 02 T 03 T 10 U 02 U 03 U 04 U 05 12_25⁎ 12_50 29_25⁎ 29_50 1000B⁎ 2000A 2000B⁎ 4000A

OS OS IS IS IS IS IS IS IS OS IS OS IS IS IS IS OS OS IS IS IS OS IS IS OS IS OS IS IS IS IS IS IS OS IS IS IS OS IS OS IS IS IS IS OS IS IS IS IS IS IS IS IS IS OS IS IS IS OS IS OS OS OS OS OS OS OS OS

⁎ 25 cm soil depth.

1997–2008 1994–2008 1993–2004 1994–2004 1993–2008 1993–2008 1993–2004 1993–2004 1993–2003 1998–2008 1998–2008 1998–2008 1992–2006 1992–2006 1993–2006 1992–2002 1997–2008 1997–2008 1990–2004 1997–2006 1997–2008 1997–2008 1988–2004 1997–2008 1997–2008 1998–2008 1996–2008 1998–2008 1987–1996 1987–2008 1987–2000 1997–2008 1990–2008 1997–2008 1997–2008 1989–2008 1998–2008 1997–2008 1991–2008 1991–2008 1990–2006 1990–2006 1990–2002 1997–2006 1997–2008 1997–2006 1997–2008 1991–2000 1991–2008 1991–2000 1991–2003 1997–2008 1997–2008 1997–2008 1997–2008 1994–2006 1993–2003 1993–2004 1994–2008 1994–2004 1994–2003 1994–2003 1994–2003 1994–2003 1992–2003 1992–2003 1992–2003 1992–2003

25 34 25 17 39 31 21 32 25 27 19 26 33 22 32 29 18 31 42 18 26 34 38 32 36 30 31 27 20 53 23 30 47 36 20 51 30 35 49 50 47 44 21 26 36 27 30 27 53 30 33 32 32 35 35 24 32 17 41 30 12 12 13 12 16 18 16 16

29 35 27 21 41 34 23 32 25 31 23 30 35 27 34 29 23 34 43 20 27 35 42 35 36 31 37 32 28 62 28 31 50 36 25 55 31 35 53 53 49 47 24 26 36 27 31 27 54 30 34 33 33 35 35 29 32 24 43 30 14 14 15 14 18 21 19 18

11.0 13.0 12.0 16.0 10.0 12.1 14.0 8.1 6.6 10.4 9.6 14.5 7.8 20.0 7.6 14.0 15.4 4.8 5.3 8.3 11.0 6.4 5.8 9.0 4.2 6.9 20.0 8.3 20.0 6.4 13.0 5.8 17.3 7.5 18.1 4.8 8.1 2.6 6.7 8.8 6.8 2.9 10.4 2.8 3.2 10.0 11.0 16.0 10.8 14.5 4.3 3.5 12.0 3.2 3.1 11.0 5.1 23.5 9.8 3.3 5.4 5.8 5.9 7.1 8.7 7.6 8.2 4.9

pH

−1

Ionic strength mmole l

5.74 4.96 5.29 6.67 5.79 5.29 5.74 6.11 5.31 5.39 5.60 5.91 5.18 4.90 5.49 5.46 5.11 4.99 4.70 5.13 4.75 4.68 4.69 4.68 5.20 5.14 6.24 4.68 4.18 4.54 4.72 4.52 4.27 4.63 4.26 4.77 4.28 4.93 4.50 4.65 5.05 4.77 4.46 4.67 4.94 4.82 4.79 4.87 4.78 5.16 4.81 5.56 5.04 5.01 5.00 5.30 4.77 6.20 4.54 5.27 4.54 4.52 4.55 4.53 4.39 4.53 4.44 4.57

0.53 1.12 0.78 1.26 0.64 0.76 1.19 0.72 0.38 1.07 0.70 0.51 1.17 1.04 0.53 0.87 0.54 0.38 1.04 0.30 0.43 0.71 0.40 0.28 0.70 1.06 0.43 0.55 1.60 0.47 0.91 0.90 2.21 0.49 1.23 0.63 1.47 0.32 0.74 0.55 0.36 0.77 0.79 0.67 0.29 0.51 1.25 0.45 0.39 0.57 0.35 0.33 0.22 0.31 0.19 0.38 0.39 0.85 1.03 0.19 0.65 0.91 0.67 0.88 1.05 0.98 1.03 0.86

−1

n+ Al

Alt mg l 0.47 1.12 0.56 0.20 0.30 0.79 0.24 0.40 0.32 0.59 0.38 0.53 0.37 1.31 0.37 0.55 0.58 0.49 1.78 0.28 0.97 1.33 1.03 0.78 0.44 1.15 0.63 1.26 1.76 1.32 1.09 2.47 5.11 1.39 3.81 0.68 4.49 0.32 1.34 1.52 0.63 0.83 1.77 1.26 0.76 0.75 1.21 0.90 0.96 0.73 0.64 0.14 0.96 0.42 0.39 0.48 0.99 0.24 2.47 0.31 2.14 2.28 1.85 1.73 2.62 2.11 2.66 2.03

−1

Aln+

Ca2+

Mg2+

Na+

K+

SO2− 4

Cl−

NO− 3

ANC

169 247 243 378 186 216 281 262 100 251 133 161 301 188 112 311 131 58 91 90 43 53 18 20 44 168 145 78 162 23 76 34 171 18 70 46 20 30 38 27 70 90 42 46 30 61 114 68 42 69 37 118 35 30 34 113 28 169 65 35 43 58 32 50 32 48 51 40

75 164 122 206 84 108 243 62 46 183 161 76 162 170 104 103 69 42 140 27 38 62 21 24 65 148 99 50 177 20 153 45 204 24 75 130 70 38 73 64 53 144 60 65 30 88 189 85 41 146 42 20 26 42 14 51 46 168 69 16 42 53 44 66 71 86 100 71

156 267 141 205 150 178 200 151 109 395 239 152 306 285 135 240 131 112 374 121 107 297 139 110 518 484 66 135 362 186 304 338 495 234 364 490 584 216 355 354 145 363 313 323 156 298 858 105 190 172 114 89 73 130 62 126 88 207 305 62 194 296 254 304 304 286 238 277

7 6 21 14 15 19 23 9 4 4 26 17 6 42 17 4 29 13 6 8 9 4 4 9 2 5 28 6 15 12 12 3 3 8 4 13 5 9 30 6 8 4 21 12 3 7 10 7 3 4 6 2 3 8 5 12 14 4 8 4 4 3 3 3 3 3 3 3

170 455 329 367 234 272 457 198 140 431 263 133 519 376 188 235 196 158 481 92 141 307 159 89 278 427 58 187 573 173 389 377 708 165 259 191 374 101 225 131 120 244 270 239 90 138 436 170 114 233 137 116 49 117 82 116 167 257 427 90 322 445 301 413 419 432 420 361

112 208 133 159 114 182 200 140 89 305 194 101 236 254 118 233 151 78 281 107 112 188 99 117 356 406 63 147 340 157 216 280 621 220 281 521 601 201 431 382 145 441 285 307 133 302 681 61 152 95 94 76 29 105 59 118 68 100 188 23 156 188 187 198 193 179 228 190

0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.2 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.3 0.5 0.4 0.2 0.1 0.1 0.1 6.4 0.1 137.2 0.1 157.1 0.1 0.2 0.1 0.1 0.1 0.4 6.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1 0.1

98 38 61 265 75 53 103 152 36 87 88 177 37 43 43 91 40 –17 –152 22 –16 –87 -62 –24 12 3 191 –48 –185 –88 9 -242 –426 –119 –227 –33 –472 –16 –127 –51 18 -72 –159 –114 –12 19 –13 26 –3 71 –47 36 59 –12 –24 48 –59 179 –172 9 –215 –243 –169 –174 –221 –186 –249 –194

−1

μeq l 1.09 1.95 1.58 0.07 1.03 1.58 1.09 0.68 1.56 1.47 1.24 0.90 1.70 2.01 1.36 1.40 1.78 1.91 2.23 1.76 2.18 2.25 2.25 2.25 1.68 1.75 0.54 2.25 2.80 2.41 2.21 2.43 2.70 2.31 2.71 2.15 2.69 1.98 2.45 2.29 1.85 2.15 2.50 2.27 1.97 2.10 2.13 2.04 2.14 1.73 2.12 1.29 1.86 1.89 1.90 1.57 2.16 0.59 2.41 1.61 1.09 1.95 1.58 0.07 1.03 1.58 1.09 0.68

18 80 32 1 13 47 10 11 20 30 18 18 23 100 19 27 42 34 143 19 79 110 87 64 27 76 12 105 175 121 82 230 503 117 391 50 448 22 119 128 44 66 168 102 55 59 96 66 78 52 48 6 69 29 27 29 78 6 218 18 200 204 164 160 246 193 251 181

1922

S. Löfgren, T. Zetterberg / Science of the Total Environment 409 (2011) 1916–1926

Table 4 Number of sites exhibiting no, negative or positive trends (p b 0.10, Seasonal Kendall, Thiels slope) in soil solution for dissolved organic carbon (DOC), protons (H+), ionic strength, − − aluminium (Al3+), sulphate (SO2− 4 ), chloride (Cl ) and nitrate (NO3 ) during the period 1987–2008 in southern Sweden.

No. of sites No trend, n Negative trend, n Min slope Max slope Positive trend, n Min slope Max slope

DOC mg l−1

H+ μeq l−1

Ionic strength mmole l−1

Alt μg l−1

−1 SO2− 4 μeq l

Cl−μeq l−1

−1 NO− 3 μeq l

ANC μeq l−1

68 32 31 − 0.5 − 1.5 5 0.11 0.75

68 37 22 − 0.1 − 3.7 9 0.2 1.6

68 29 33 − 0.004 − 0.13 6 0.001 0.08

68 36 23 − 13 − 179 9 8 176

68 18 49 −2 − 42 1 3

68 45 13 −1 − 36 10 6 48

68 65 1 − 0.03

68 33 3 −2 − 33 32 2 21

number of sites showing decreased, increased or no DOC concentration trends are shown. At the sites showing recovery from acidification, decreased DOC concentrations are the prevailing trend. With respect to increases in pH, 59% (n = 13) of the sites exhibit a negative DOC trend, which is the opposite direction to what to be expected from pure pH increases without simultaneous changes in ionic strength and aluminium content. At such conditions, the net charge and solubility of the organic matter should increase (c.f. Löfgren et al., 2010 and references therein). However, only one site out of 22 exhibited the expected, positive DOC trend at the same time as pH increased (Table 6). Löfgren et al. (2010), Hruska et al. (2009) and Wu et al. (2010) found the same poor relation between increasing pH and DOC trends in soil solution. Thus, it must be concluded that the recovery from acidification, with respect to increased pH, seems to have had a very low impact on the DOC concentrations in soil water. Instead, other chemical factors must be responsible for the decreasing DOC trends. The decreased DOC concentrations at sites with statistically significant reduced ionic strength and Alt concentrations are in agreement with what geochemical modelling exercises have revealed on soil solutions from the region and with similar chemistry (Löfgren et al., 2010). Reduced ionic strength and increased soil aluminium pools, coupled to a lower Al solubility in soil water, seem to increase the coagulation of organic matter and thereby reduce the DOC concentrations (Löfgren et al., 2010). In the Löfgren et al. (2010) study, the concentration of DOC, pH, ionic strength and Alt varied

2 2 20

between 3.7–24.2 mg DOC l−1, 4.3–5.4, 0.14–2.67 mmole l−1 and 0.39–5.1 mg Alt l−1, respectively. These intervals are very similar to those found in the present study (Table 3). Thus, it is not surprising that the DOC concentrations decreased at 57% (n = 20) and 70% (n = 16) of the sites with reduced ionic strength and aluminium concentrations and that only two sites with reduced ionic strength showed the opposite trend direction (Table 6). No site exhibited a positive DOC trend at sites with decreased Al concentrations. Hence, both this study and the work by Löfgren et al. (2010) show that the reduced ionic strength and changed aluminium chemistry, mainly coupled to the reduced sulphur deposition, appear to be responsible for the decreased DOC concentrations in the soil solution. In contrast to our results, Hruska et al. (2009) found a negative relation between ionic strength and DOC in soil water at two sites in the Czech Republic. At the acidic Lysina catchment, pH was in the range 3.2–3.8 and the ionic strength varied between 0.2 and 1.4 mmole l−1. At the well-buffered site Pluhov Bor, the same ranges were 5–6 for pH and 1–4 mmole l−1 for the ionic strength (Hruska et al., 2009). In our study, the median pH and ionic strength varied in the ranges 4.2–6.7 and 0.2–2.5 mmole l−1, respectively (Table 3). Hence, except for the low pH at Lysina, the two studies cover more or less the same ranges. Therefore, we do not think that the principle explanation for the different DOC trend directions is to be found in pH or ionic strength. Another and maybe more important difference between the Hruska et al. (2009) and our study is that the soil solution was

− + Fig. 2. Trends in the soil water concentrations of sulphate (SO2− 4 ), chloride (Cl ), acid neutralizing capacity (ANC), dissolved organic carbon (DOC), protons (H ), ionic strength (IS) and total aluminium (Alt) at the 68 sampling sites in southern Sweden. Blue dot = decreasing concentrations, Red dot = increasing concentrations, Grey dot = no trend.

S. Löfgren, T. Zetterberg / Science of the Total Environment 409 (2011) 1916–1926

1923

Fig. 2 (continued).

collected at different soil depths. At Lysina, the lysimeters were placed under the forest floor of a podzol and at Pluhov Bor in the top-mineral soil of a brown earth, while the samplers in our study mainly were placed at 50 cm soil depth (25 cm at 4 sites) in podzols. Data from the Swedish forest soil survey (1983–87) show that the median humus layer thickness in podzols (semi-mature and mature coniferous forests on sorted sediments and glacial till) is 8 cm and that less than 1% have a forest floor thickness N30 cm (Löfgren et al., 2009a; Löfgren et al., 2008). Thus, the soil solution data in our study represent lower B-horizon conditions and one would expect most of the DOC trapped in the upper B-horizon due to the podzolization process (Fröberg et al., 2006; Lundström et al., 2000). This is not the case for the Hruska et al. (2009) data, which is collected above this horizon.

Similarily, the UK data from shallow soils also indicate increasing DOC concentrations, but during a much shorter time period (Buckingham et al., 2008). The Norwegian results (Wu et al., 2010) from 5, 15 and 40 cm soil depth, roughly corresponding to the base of the organic horizon and the upper and lower parts of the mineral soil, support our hypothesis by showing statistically significant decreasing DOC trends primarily in the top mineral soil (15 cm). It is well known that the quantitative importance of chemical and biological processes determining the DOC concentration is significantly different above and below the B-horizon (Fröberg et al., 2006). Hruska et al. (2009) also states that changes in the production of DOC in the forest floor may be important for the observed trends. They also suggest changes in ionic strength in deep mineral soils as a driving force for the observed DOC

1924

S. Löfgren, T. Zetterberg / Science of the Total Environment 409 (2011) 1916–1926

Table 5 Data period (nyr = number of years) and trend rates (p b 0.10, Seasonal Kendall, Thiels slope) for dissolved organic carbon (DOC), protons (H+), pH, ionic strength and aluminium (Alt) in soil solution at 68 sites in southern Sweden monitored during the period 1987–2008. − = negative trend and + = positive trend. Site

A01 A05 A21 A24 A35 A40 A44 A54 A90 A92 D11 D14 D51 D52 E02 E03 E21 E22 F09 F21 F22 F23 G09 G21 G22 H03 H21 H22 K01 K03 K10 K13 L05 L07 M13 N01 N17 O01 O05 O35 P02 P52 P70 P92 P93 P94 R09 S01 S05 S15 S16 S22 S23 T02 T03 T10 U02 U03 U04 U05 12_25# 12_50 29_25# 29_50 1000B# 2000A 2000B# 4000A

nyr

12 15 12 11 16 16 12 12 11 11 11 11 15 15 14 11 12 12 15 10 12 12 17 12 12 11 13 11 10 22 14 12 19 12 12 20 11 12 18 18 17 17 13 10 12 10 12 10 18 10 12 12 12 12 12 13 11 12 15 11 10 10 10 10 12 12 12 12

⁎ p b 0.10. ⁎⁎ p b 0.05. ⁎⁎⁎ p b 0.01. # 25 cm soil depth.

DOC

H+

pH

Ionic strength

Alt

mg l− 1 year− 1

μeq l− 1 year− 1

year− 1

mmole l− 1 year− 1

mg l− 1 year− 1

− 0.9⁎⁎

+ 0.04⁎⁎

− 0.13⁎⁎

+ 0.18⁎ + 0.44⁎⁎⁎

− 0.38⁎

− 0.1⁎⁎ − 0.4⁎ + 0.4⁎

+ 0.05⁎⁎ + 0.03⁎ − 0.05⁎

+ 0.2⁎

− 0.07⁎

− 0.4⁎⁎⁎ − 0.12⁎ − 0.47⁎⁎⁎ − 0.47⁎⁎⁎ − 0,5⁎ − 0.17⁎⁎⁎

− 0.15⁎⁎⁎ − 0.35⁎⁎⁎ − 1.47⁎⁎

− 0.42⁎ − 1⁎⁎ − 0.25⁎⁎⁎ − 0.71⁎⁎⁎ + 0.27⁎⁎ − 0.34⁎⁎⁎ − 0.24⁎⁎⁎ − 0.41⁎⁎ − 0.24⁎⁎ − 0.25⁎⁎⁎ + 0.108⁎

− 0.40⁎ − 0.30⁎ − 0.30⁎ − 0.60⁎⁎⁎

− 13⁎ + 18⁎ − 20⁎⁎

+ 0.056⁎ + 0.029⁎ − 0.014⁎

+ 12⁎

− 0.040⁎⁎ − 0.019⁎⁎ − 0.088⁎⁎⁎ − 0.5⁎

+ 0.02⁎

− 0.2⁎ − 0.15⁎⁎

− 0.024⁎ − 0.017⁎

− 0.6⁎⁎ + 0.75⁎⁎ − 0.34⁎⁎⁎ − 0.3⁎⁎⁎ − 0.5⁎ − 0.71⁎⁎

− 15⁎⁎

+ 0.5⁎⁎

− 0.03⁎⁎

− 2.7⁎⁎⁎ − 1.3⁎

+ 0.04⁎⁎⁎ + 0.03⁎

− 0.8⁎⁎

+ 0.01⁎⁎

− 0.8⁎⁎⁎ + 1.6⁎⁎⁎

+ 0.03⁎⁎⁎ − 0.02⁎⁎⁎

− 0.5⁎⁎⁎ − 0.4⁎ − 3.7⁎

+ 0.02⁎⁎⁎ + 0.01⁎ + 0.05⁎

− 0.3⁎⁎ + 0.6⁎ − 1.0⁎⁎⁎ + 0.9⁎⁎⁎ − 0.6⁎⁎⁎ − 0.8⁎ + 0.5⁎ + 0.2⁎⁎

+ 0.01⁎⁎ − 0.02⁎ + 0.02⁎⁎⁎ − 0.03⁎⁎⁎ + 0.02⁎⁎⁎ + 0.05⁎ − 0.01⁎ − 0.04⁎⁎

+ 0.5⁎⁎⁎

− 0.02⁎⁎⁎

− 0.2⁎⁎ − 0.8⁎⁎⁎ − 0.1⁎⁎

+ 0.08⁎⁎ + 0.01⁎⁎⁎ + 0.01⁎⁎

− 0.7⁎⁎ − 0.6⁎ − 1.8⁎⁎

+ 0.01⁎⁎ + 0.01⁎ + 0.03⁎⁎

− 1.2⁎⁎

+ 0.01⁎⁎

− 0.010⁎⁎⁎ − 0.038⁎⁎⁎

− 0.010⁎ − 0.032⁎⁎⁎ + 0.033⁎⁎⁎ − 0.011⁎ − 0.037⁎⁎⁎ − 0.115⁎⁎ − 0.037⁎⁎⁎ − 0.129⁎⁎⁎ − 0.094⁎⁎⁎ − 0.022⁎ + 0.084⁎⁎⁎ − 0.059⁎⁎⁎ + 0.001⁎ + 0.011⁎ − 0.015⁎ − 0.018⁎⁎⁎ − 0.030⁎⁎⁎ − 0.111⁎⁎⁎ − 0.008⁎⁎

− 0.015⁎⁎⁎ − 0.032⁎⁎ − 0.004⁎

− 0.062⁎⁎⁎

+ 19⁎⁎

− 79⁎⁎⁎ + 47⁎ − 72⁎⁎⁎ − 124⁎ − 113⁎⁎⁎ − 83⁎⁎ + 176⁎⁎ − 45⁎⁎⁎ + 45⁎⁎ − 31⁎⁎ − 29⁎⁎ − 21⁎⁎⁎ − 28⁎⁎ − 179⁎⁎⁎ − 54⁎⁎ − 40⁎⁎⁎ + 13⁎ − 37⁎ − 20⁎⁎⁎ + 29⁎⁎ + 8⁎⁎⁎ − 16⁎⁎⁎ − 34⁎⁎

− 0.005⁎⁎⁎

− 0.069⁎⁎ − 0.081⁎⁎⁎ − 0.041⁎⁎⁎ − 0.053⁎⁎ − 0.030⁎⁎

− 119⁎⁎⁎ − 60⁎ − 106⁎⁎

S. Löfgren, T. Zetterberg / Science of the Total Environment 409 (2011) 1916–1926

DOC mg l-1

18 16 14 12 10 8 6 4 2 0

1925

pH

6 5.5 5 4.5 4 3.5

2006

2004

2002

2000

1998

1996

1994

1992

1990

2006

2004

2002

2000

1998

1996

1994

1992

1990

3

Alt mg l-1

Ionic strength mmol l-1 1.6

2.5

1.4 2

1.2 1

1.5

0.8 1

0.6 0.4

0.5

0.2 2006

2004

2002

2000

1998

1996

1994

1992

2006

2004

2002

2000

1998

1996

1994

1992

1990

1990

0

0

Fig. 3. Time series (1990–2005) of the soil water concentrations of dissolved organic carbon (DOC), pH, ionic strength and total aluminium (Alt) at the P52 site.

trends (increasing) in stream water. However, according to our results and Löfgren et al. (2010), decreased ionic strength would mainly increase the DOC coagulation in mineral soils in the B horizon. Opposite to the often found increased DOC concentrations in surface waters in the northern hemisphere (Erlandsson et al., 2008; Evans et al., 2005; Monteith et al., 2007), the DOC concentrations in soil water decreased or showed no trend below the B-horizon in recharge areas. The generally small increases in pH (median b0.3 pH units) during the investigation period seem to be counterbalanced by the reduced ionic strength and diminished Alt concentrations, increasing the organic matter coagulation. Hence, in contrast to the conclusion for surface waters by Clark et al. (2010), the solubility of organic matter seems to decrease in uphill soils, as a result of the acidification recovery. The low hydraulic connectivity between soil solution and groundwater in recharge areas and streams (Seibert et al., 2009), causing large modifications of the pH, Al and DOC chemistry in groundwater along the hillslopes (Löfgren and Cory, 2010), provide difficulties to apply results from mineral soils in recharge areas (this study and e.g. Wu et al., 2010) for rejecting the DOC acidification recovery hypothesis by Clark et al. (2010). However, in combination with the DOC chemistry and trends in soil solution in riparian soils (Löfgren et al., 2010) it seems unlikely that recovery from acidification could be the main driver for the increased DOC concentrations in surface waters. This conclusion is valid also for soil water studies above the B-horizon, where the hydraulic connectivity would be even

Table 6 The number of sites with signs of recovery from acidification in terms of statistically significant (p b 0.10) positive trends in pH, negative trends in ionic strength and decreased aluminium concentrations. For each chemical factor, the number of sites (nDOC) showing decreased, increased or no DOC concentration trends are summarised.

pH increase, n = 22 Ionic strength decrease, n = 35 Aluminium decrease, n = 23

nDOC — decrease

nDOC — no trend

nDOC — increase

13 20

8 13

1 2

16

7

0

lower and DOC transport to the streams feasible only through lateral flow during extreme runoff events. Finally, the hydraulic connectivity put focus on where to search for the processes determining the DOC trends in surface waters. Our results support the theory that processes in riparian zones and peat lands rather than dry soils uphill govern the stream water DOC variation (Köhler et al., 2009). Hence, the processes controlling surface water DOC trends should be searched for in such organic rich soils with high connectivity to streams and lakes. In these areas, variations in the meteorological and hydrological conditions have a great impact on the DOC chemistry and biology, complicating the studies of DOC loadings to surface waters. 5. Conclusions The main aim with this study was to identify potential drivers for the DOC solubility in soil solution by comparing trends in DOC concentrations with observed changes in pH, ionic strength and concentrations of Aln+. As in earlier Swedish and Norwegian studies, the DOC concentrations in soil water decreased or showed no trend. The generally small increases in pH (median b0.3 pH units) during the investigation period seem to be counterbalanced by the reduced ionic strength and diminished Al concentrations, increasing the organic matter coagulation. Hence, opposite to the conclusion for surface waters, the solubility of organic matter seems to decrease in uphill soils, as a result of the acidification recovery. The results support the theory that processes in riparian zones and peat lands rather than dry soils uphill govern the stream water DOC variations. Hence, the processes controlling the surface water DOC trends should be searched for in such organic rich soils with high connectivity to streams and lakes. Acknowledgement This research has been supported by the Swedish Environmental Protection Agency, the Swedish Forest Agency, the Swedish University of Agricultural Sciences, SLU and the Foundation for the Swedish Environmental Research Institute, IVL. The Swedish Throughfall Monitoring Network is run by IVL Swedish Environmental Research

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Institute and financed by regional air quality protection associations, county administrative boards and the Swedish Environmental Protection Agency. References Buckingham S, Tipping E, Hamilton-Taylor J. Concentrations and fluxes of dissolved organic carbon in UK topsoils. Sci Total Environ 2008;407:460–70. Clark JM, Bottrell SH, Evans CD, Monteith DT, Bartlett R, Rose R, et al. The importance of the relationship between scale and process in understanding long-term DOC dynamics. Sci Total Environ 2010;408:2768–75. Dawson JJC, Malcolm IA, Middlemas SJ, Tetzlaff D, Soulsby C. Is the composition of dissolved organic carbon changing in upland acidic streams? Environ Sci Technol 2009;43:7748–53. Driscoll CT. A procedure for the fractionation of aqueous aluminum in dilute acidic waters. Int J Environ an Ch 1984;16:267–83. Erlandsson M, Buffam I, Fölster J, Laudon H, Temnerud J, Weyhenmeyer GA, et al. Thirtyfive years of synchrony in the organic matter concentrations of Swedish rivers explained by variation in flow and sulphate. Global Change Biol 2008;14:1191–8. Evans CD, Monteith DT, Cooper DM. Long-term increases in surface water dissolved organic carbon: observations, possible causes and environmental impacts. Environ Pollut 2005;137:55–71. Fölster J, Bringmark L, Lundin L. Temporal and spatial variations in soilwater chemistry at three acid forest sites. Water Air Soil Pollut 2003;146:171–95. Fröberg M, Berggren D, Bergkvist B, Bryant C, Mulder J. Concentration and fluxes of dissolved organic carbon (DOC) in three norway spruce stands along a climatic gradient in sweden. Biogeochemistry 2006;77:1-23. Haaland S, Hongve D, Laudon H, Riise G, Vogt RD. Quantifying the drivers of the increasing colored organic matter in boreal surface waters. Am Chem Soc 2010;44:2975–80. Hallgren-Larsson E, Knulst J, Malm G, Westling O. Deposition of acidifying compounds in Sweden. Water Air Soil Poll 1995;85:2271–6. Helsel DR, Hirsch RM. Statistical measures in water research. Amsterdam: Elsevier Science B.V., 1992. Hirsch RM, Slack JR. A nonparametric trend test for seasonal data with serial dependence. Water Resour Res 1984;20:727–32. Hruska J, Kram P, Mcdowell WH, Oulehle F. Increased dissolved organic carbon (DOC) in Central European streams is driven by reductions in ionic strength rather than climate change or decreasing acidity. Environ Sci Technol 2009;43:4320–6. Karltun E, Stendahl J, Lundin L. Acid-base status and changes in Swedish forest soils. KSLAT 2003;142:31–6. Köhler SJ, Buffam I, Seibert J, Bishop KH, Laudon H. Dynamics of stream water TOC concentrations in a boreal headwater catchment: controlling factors and implications for climate scenarios. J Hydrol 2009;373:44–56. Laudon H, Hedtjärn J, Schelker J, Bishop K, Sörensen R, Ågren A. Response of dissolved organic carbon following forest harvesting in a boreal forest. Ambio 2009;38: 381–6. Löfgren S, Cory N. Groundwater Al dynamics in boreal hillslopes at three integrated monitoring sites along a sulphur deposition gradient in Sweden. J Hydrol 2010;380: 289–97. Löfgren S, Bringmark L, Aastrup M, Hultberg H, Kindbom K, Kvarnäs H. Sulphur balances and dynamics in three forested catchments in Sweden. Water Air Soil Poll 2001;130:631–6. Löfgren S, Zetterberg T, Larsson P-E, Cory N, Malin K, Kronnäs V, et al. Skogsmarkskalkningens effekter på kemin i mark, grundvatten och ytvatten i SKOKAL-

områdena 16 år efter behandling (The large-scale effects of forest soil liming on the chemistry in soil, groundwater and surface water at the SKOKAL plots 16 years after treatment). Swedish Forest Agency Report. 16:2008, Jönköping; 2008. p. 124. (In Swedish). Löfgren S, Cory N, Zetterberg T, Larsson P-E, Kronnäs V. The long-term effects of catchment liming and reduced sulphur deposition on forest soils and runoff chemistry in southwest Sweden. For Ecol Man 2009a;258:567–78. Löfgren S, Ring E, von Brömssen C, Sörensen R, Högbom L. Short-term effects of clearcutting on the water chemistry of two boreal streams in Northern Sweden: a paired catchment study. Ambio 2009b;38:347–56. Löfgren S, Gustafsson JP, Bringmark L. Decreasing DOC trends in soil solution along the hill slopes at two IM sites in southern Sweden — geochemical modeling of organic matter solubility during acidification recovery. Sci Total Environ 2010;409:201–10. Loftis JC, Taylor CH, Chapman PL. Multivariate tests for trend in water-quality. Water Resour Res 1991;27:1419–29. Lundström US, van Breemen N, Bain D. The podzolization process. A review. Geoderma 2000;94:91-107. Monteith DT, Stoddard JL, Evans CD, de Wit HA, Forsius M, Hogasen T, et al. Dissolved organic carbon trends resulting from changes in atmospheric deposition chemistry. Nature 2007;450:537-U9. Reuss JO, Johnson DW. Acid deposition and acidification of soils and waters. Ecological studies, vol 59. Springer-Verlag; 1986. Sarkkola S, Koivusalo H, Lauren A, Kortelainen P, Mattsson T, Palviainen M, et al. Trends in hydrometeorological conditions and stream water organic carbon in boreal forested catchments. Sci Total Environ 2009;408:92-101. Seibert J, Grabs T, Kohler S, Laudon H, Winterdahl M, Bishop K. Linking soil- and streamwater chemistry based on a riparian flow-concentration integration model. Hydrol Earth Syst Sc 2009;13:2287–97. Skjelkvåle BL, Stoddard JL, Jeffries DS, Torseth K, Hogasen T, Bowman J, et al. Regional scale evidence for improvements in surface water chemistry 1990–2001. Environ Pollut 2005;137:165–76. Thienemann A. Die Binnengewässer Mitteleuropas. Eine limnologische Einführung, vol 1. Stuttgart: E. Schweizerbart Science Publishers; 1925. Weng L, Temminghoff EJM, van Riemsdijk WH. Interpretation of humic acid coagulation and soluble soil organic matter using a calculated electrostatic potential. Eur J Soil Sci 2002;53:575–87. Westling O, Lövblad G. Deposition trends in Sweden. In: Warfvinge P, Bertills U, editors. The Environmental Recovery from Acidification. ReportStockholm: Swedish Environmental Protection Agency; 2000. Wu YJ, Clarke N, Mulder J. Dissolved organic carbon concentrations in throughfall and soil waters at level ii monitoring plots in Norway: short- and long-term variations. Water Air Soil Poll 2010;205:273–88. www.krondroppsnatet.ivl.se 2011-02-03 (in Swedish). www.slu.se/vatten-miljo/vattenanalyser 2011-02-03 (in Swedish). Yallop AR, Clutterbuck B. Land management as a factor controlling dissolved organic carbon release from upland peat soils 1: spatial variation in DOC productivity. Sci Total Environ 2009;407:3803–13. Zetterberg T, Akselsson C, Westling O. 1. Markvattenkemiska effekter vid spridning av aska på skogsmark. Slutrapport från ett 10-årigt försök (Soil water effects of ash application on forest soils. Final report from a 10 years study). IVL Rapport; 2006a. B1658 (In Swedish). Zetterberg T, Akselsson C, Westling O. 2. Markvattenkemiska effekter vid spridning av kalk på skogsmark. Slutrapport från ett 12-årigt försök. (Soil water effects of forest soil liming. Final report from a 12 years study). IVL Rapport; 2006b. B1652 (In Swedish).