Long Term Effects of Forest Liming on the Acid-Base Budget

: In Rhineland-Palatinate (Germany), a high percentage of the forest area is located on poor soils with low buffering capacity. Extensive liming applications were performed to compensate for the negative consequences of acid deposition. In 1988, three experimental sites with untreated control plots and different liming treatments were established in coniferous stands to investigate the effectiveness of liming on acidiﬁcation and its effect on forest ecosystems. Measuring deposition and seepage waters for 24 years allowed for calculating long-term acid-base budgets. The original approach was expanded by data from a detailed sampling of the forest stand and mineral weathering rates. Without liming, the acid load exceeded the buffer capacity by base cation release from silicate weathering during the whole observation period. As a result, there was a high release of aluminum. After liming seepage water output of organic anions, nitrate and sulfate increased in some cases, leading to a higher acid load. However, the carbonates of dolomitic limestone compensated for a higher acid load, resulting in less aluminum released compared to the control plots. Until sulfate output by seepage water declines and nitrogen emissions are reduced, liming and restricted biomass harvesting are required for forest stands on base poor soils to prevent further acidiﬁcation, decline of nutrient stocks, and the destruction of clay minerals.


Introduction
Since the late 19th century, an increased acid input from anthropogenic activities was observed in Europe and North America [1]. Forest ecosystems were especially affected when their position was exposed and because of their large intercepting canopy surface [2]. Since the 1980s, acid atmospheric deposition has been decreasing, mainly due to reduced sulfur dioxide emissions [3][4][5]. However, the input of nitrogen components still remains at a high level (cf. [6]) and is currently the main source for the acid load entering the forest ecosystems [7,8].
In addition, centuries of intensive litter and timber harvesting contributed to large scale soil acidification [9,10]. In Rhineland-Palatinate, which is located in southwestern Germany (Figure 1), a high percentage of forested areas are located on base-poor soils with small nutrient reserves and low buffer capacity against acidity [11]. As a result, mobilization of aluminum and heavy metals, reduction of base cation reserves, and destabilization of clay minerals could, and still can, be observed for many forest sites [12][13][14][15]. To compensate for the negative consequences of the acid deposition, extensive liming actions with dolomitic lime were performed at an early stage [16,17]. The application of 3 to 4 tons per hectare every 10 years was recommended [2,18].
In 1988, three experimental sites ( Figure 1) with different liming treatments were established. The base-poor forest sites were treated once to evaluate the forest management practice of liming and to investigate its effectiveness and impacts on forest ecosystems [19]. ecosystems [19]. Long term input-output element and acid-base budgets were calculated (cf. [14]) to characterize the forest ecosystems of the three experimental sites by the different processes causing the fluxes of acidity and to quantify soil acidification without and with different dosages of dolomitic limestone. Therefore, we investigated if the mobilization of aluminum could be observed in the control plots without liming, which would indicate that proton production processes exceeded the proton consumption by reaction associated with Mb cations (=Ca, K, Mg, Na). In this case, proton consumption is carried out partially by the weathering of Al, Mn and Fe (Ma cations) oxides causing the destabilization of clay minerals [20]. Further, important questions include did the liming treatments counter the effects of acidification in the long term without the negative consequences of nitrate mobilization? These questions will be answered by input-output budgeting.

Sites
Each of the three experimental sites, Adenau (AD), Idar-Oberstein (IO), and Hochspeyer (HS) ( Table 1), was subjected to five liming treatments, ranging from 3 to 15 t ha −1 dolomitic limestone with untreated control plots ( Table 2). The area of each liming treatment is 2000 m 2 separated into two subplots of 1000 m 2 ( Figure 1). The control treatment has three subplots of 2125 m 2 each. The experimental sites are fenced, and managed forest stands are thinned regularly. Lime and fertilizers were spread by hand in December 1988.  Table 2. More information about the experimental sites is documented in Table 1.  Table 2. More information about the experimental sites is documented in Table 1.  Since the establishment of the study, the sampling of seepage water occurred at depths of 60 cm and 10 cm using suction cups and directly below the humus layer by funnel lysimeters. Four suction cups per depth and five funnel lysimeters were installed on each subplot. Six continuously open bulk samplers with a total collection area of 1885 cm 2 were used to collect the bulk deposition at a nearby clearing and the throughfall on two of the control subplots. The water samples were collected every two weeks and kept in cold storage. For each subplot, the collected water was analyzed once every three months as a mixed sample for each of the different sampling types. The analyses were performed according to the methods of the Handbuch Forstliche Analytik [21].

Element Fluxes
The soil water fluxes for each experimental site were calculated by a calibrated COUP-MODEL, as described in Karl et al. [23]. To calculate the element fluxes, the element concentrations in the seepage water are multiplied with the soil water fluxes, for each depth (cf. [24,25]). The total deposition (TD) was calculated for each experimental site using the bulk deposition and throughfall measurements as inputs into the canopy budget model after Ulrich [26] and Draaijers and Erisman [27]. For the total annual N deposition, the maximum from both canopy budget models plus the input of organic N of the bulk deposition was used, because total N deposition derived from canopy budget models is typically underestimated [8,28].
Additionally, weathering rates, calculated by PROFILE (cf. [29]), were used to derive the proton consumption from base cation release for each experimental site. The updated PROFILE version 4.4 was used, which includes typical minerals found in German soils [30]. A quantitative mineral analysis was performed in 1997 and the surface area of the mineral soil was calculated based on the soil texture, the coarse soil, and the dry bulk density (DBD) after Becker [30]. The long term mean air temperature was used for the soil temperature (cf. [31]). The plant available soil water content (water content reduced by the non-plant available water content below the permanent wilting point) was calculated by the COUP-MODEL and used for soil moisture input to the PROFILE model. The influence of CO 2 was removed from the calculations for all minerals by setting the "Rate Constant_CO2" to 30, because of its weak effect on mineral dissolution (cf. [32][33][34]).
The acid load by biomass increment was also calculated. The single tree-based stand simulator SILVA [35], which was adjusted with allometric relations for Rhineland-Palatinate [36,37], was used to calculate the above ground biomass of the different tree compartments for the years 1988 and 2011 (IO, HS), respectively 1988 and 2013 (AD) for each subplot. As input data, the diameter at breast height (DBH) was measured for each tree and the height for every third tree on the experimental site in winter 1988/89 and 2011/12 (IO, HS) respectively 1988/89 and 2013/14 (AD).
In the winter of 2011/12, nine trees of the control plots (treatment 0) and six trees per liming treatment 1 (3 t ha −1 ), 3 (3 t ha −1 +P), and 8 (15 t ha −1 +P) were felled on each experimental site to obtain information about the incorporated nutrients. These trees were divided into needles, twigs, branches, bark, and wood according to the method described by Pretzsch et al. [37]. Additionally, the wood of the Scots pine trees was separated into sapwood and heartwood. A mixed sample of each compartment per tree was analyzed separately. Under the assumption, the element concentrations gained of the trees from the control plot were identical to the element concentrations at the beginning of the experiment, they were used to calculate the above ground element stocks for all liming treatments and the control at the start of the liming trial in 1988. The above ground element stocks of the years 2011 (IO, HS) respectively 2013 (AD) were calculated by the element concentrations of the trees felled from the according liming treatments. The element concentrations of the liming treatments 6 and 7 were interpolated by linear regression, because the needle and litterfall samples of all liming treatments indicate a significant correlation between liming dosage and element concentration [22]. The differences of the element stocks between these two dates represent the amount of elements incorporated since the beginning of the liming trial. The results were adjusted to account for the period of the other element fluxes by calculating the mean annual incorporation and multiplying it by 24 years.

Calculation of Acid-Base Budgets
The input-output budgets for the forest soil were calculated based on element fluxes (cf. [14,26]) on a yearly basis for 24 years, from 1989 to 2012. The inputs to the soilinternal element cycle are total deposition and elements released by mineral weathering. The element loss by seepage water is an output from the forest soil. Lime and fertilizers are treated as part of the soil and not as an additional input.
The acid load based on the budgets of NH 4 + , H + , Mn 2+ , Al 3+ , Fe 3+ , SO 4 2− , NO 3 − , organic anions (Org − ), Ca 2+ , K + , Mg 2+ , and Na + was calculated for each subplot after Ulrich [26,38]. Dissolved organic carbon (DOC) was measured in the water samples and converted to Org − , as described in Mosello et al. [39] with the updated conversion factor of ICP Forests [40]. P was not included because of uncertainties in the input-output budgeting. The element fluxes at 60 cm depth were taken as an output from the ecosystem and subtracted from the accordant element input by TD.
In the original approach by Ulrich [26], mineral weathering and the acid load by biomass increment are not calculated separately but instead included in the element budgets. Positive budgets of NH 4 + , H + , Mn 2+ , Al 3+ , or Fe 3+ account for a net input of acidity into the ecosystem because these ions count as potential proton donors [14]. Moreover, positive budgets of Ca 2+ , K + , Mg 2+ , and Na + add to the acid load, because it is assumed that these M b cations are incorporated into the biomass increment or are bound to exchange sites of the mineral soil and releasing M a cations that lead to proton production. Negative budgets of SO 4 2− , NO 3 − or Org − represent the dissolution of aluminum sulfates, nitrification, or the dissociation of dissolved organic acids and contribute also to the acid load. Opposing results of the element budgets lead to proton consumption by acid base reactions like dissolution of M b and M a oxides, cation exchange, mineralization, or the formation of aluminum sulfates [26].
Because of the decision to treat the added lime and fertilizer as part of the soil, the M b cation budgets of the liming treatments are more negative in comparison to the control plots. Lime and fertilizer could also be treated as an additional input, which would lead to positive M b cation budgets and therefore adding to the acid load for all liming treatments. This acid load originating from the M b cation input by liming would have to be balanced by protonation of HCO 3 − , which would complicate the calculations as well as the presented tables, without adding additional insights.

Modification of the Original Approach
In our modified approach, we use the biomass sampling of forest stand and the PRO-FILE calculations to get a detailed look at the budget of M b cations in the original calculation.
M b cations are also taken up by the forest stand and are incorporated into the biomass, which leads to proton production. For the calculation of the acid load, the incorporation of elements in the aboveground biomass (net element uptake by the forest stand that is needed for long-term growth) is taken into account. The element uptake to supply roots, leaves, and needles, which is not removed by harvesting, or returned to the forest floor by fine root turnover or litterfall, is part of the ecosystem-internal element cycle and was not included in the input-output budgets. The cation excess equals the proton production by biomass increment [26] and was calculated based on the analyzed elements: In this calculation, the influence of the utilized N form is not taken into account, because of the uncertainty of their availability to and uptake by the forest stand. It is assumed that both N forms are taken up in equal shares. A higher NH 4 + uptake would increase the acid load whereas a higher NO 3 − uptake would lead to a lower acid load by biomass increment [41].
To calculate the M b cation exchange and dissolution of calcium and magnesium carbonates included in the dolomite, the amount of M b cations incorporated into the biomass and lost by seepage water output are subtracted from M b cation release by mineral weathering and deposition input. For the control plots, this difference represents the proton consumption or production by M b cation exchange processes. On the liming trials, the dissolution of carbonates is also included in this difference.
Differences between the budgets of the subplots within an experimental site are the result of differences in element loss by seepage, as well as the incorporation of elements in the forest stand, which were measured separately for each subplot. Deposition and mineral weathering were assumed to be identical for all subplots of an experimental site.

Site Characteristics
We use the original approach to characterize the three sites because the complex processes are summarized to a greater extent. The acid load and the acid-base reactions, based on the element budgets of the control plots, show clear differences in their composition between the three experimental sites (Table 3). At AD, the main source of proton production is the N budget because of high NH 4 + input. In some years, the NO 3 output exceeds the NO 3 − input, also leading to proton production. In IO, the input of NH 4 + is lower and almost all the N is retained in the ecosystem so that the combined N budget of NH 4 + and NO 3 − still adds to the proton consumption. Instead of N, the release of stored SO 4 2− and the retention of H + contribute primarily to the acid load of this forest site. In HS, the main source of proton production is the loss of organic acids in combination with the input of NH 4 + . Like in IO, almost all NO 3 − is retained in the ecosystem, noticeably contributing to the proton consumption.
The proton consumption for all control plots of the three study areas occurs mostly by M a cation exchange or weathering, especially of aluminum. Except for subplot 2 in AD, the accumulation of NO 3 − contributes noticeably to the proton consumption. In IO and HS, higher proton consumption is calculated compared to the proton production (Difference unequal zero). This could be caused by an underestimation of the flux of organic acids (cf. [26]). Additionally, natural variation inside the experimental plots may contribute to the errors.

Effect of Liming Treatments
The liming treatments in AD show a higher acid load compared to the control plots. The low-dose as well as the high-dose liming treatments increased the dissolution of sulfates and the nitrification, leading to a greater loss of S and N by seepage water flux [42]. Especially the lower retention of NO 3 − reduces the proton consumption by the acid-base reactions of the N budget in comparison to the control plot. The high acid load of treatment 7 is caused by the additional application of sulfur bound K and Mg fertilizers (cf. Table 2), which leads to a high sulfate flux by seepage water accompanied by cation loss.
The forest ecosystem in IO shows a different reaction to the liming treatment. Acid load does not increase on the different liming treatments in general. There are plots of the liming treatments with similar, higher, or lower acid load compared to the control plots. In HS, some liming treatments show higher acid load due to higher flux of organic acids with the seepage water, though the acid load in HS for all treatments is lower than even the acid load of the control plots in AD and IO.
For all three experimental sites, the bigger part of the acid load on the control plots is compensated by M a cation release. Although the acid load increases for most liming treatments, the proportion of proton consumption by M a cations decreases with increasing dosage for all study areas. In AD and HS, the absolute portion [keq ha −1 ] remains on a similar level to the control plots and in IO both absolute portion and relative proportions decrease. This shows clearly that the additional acid load of reactions triggered by the liming is compensated by the dissolution of calcium and magnesium carbonates included in the dolomite.
The contribution of the different element budgets to H + buffering or production is shown more clearly when proton production and consumption cations (which are shown separately for the control plots in Table 3) are summed up (Table 4). In this table, we also included the additional data of mineral weathering and biomass production to examine the effects of the liming treatments and the involved processes on the M b budget in more detail.
For the liming treatments, the acid load through biomass increment increases compared to the control plot because of higher incorporation of Ca and Mg in all biomass compartments and because of increase in biomass increment [22]. With regard to this calculation, it is assumed that NH 4 + and NO 3 − are taken up with the same proportion of 50%. In HS and AD an uptake of N only as NH 4 + could almost double the acid load compared to a proportion of 50% NH 4 + . In the case of IO an uptake of N only in the form of NO 3 − would not only lead to a lower acid load, but result in proton consumption by biomass increment for the liming treatments ( Figure 2).
The M b cation input by mineral weathering and deposition is on all control plots not high enough to compensate the loss by seepage water and uptake by the forest stand. This indicates that whole-tree harvesting is not nutrient sustainable for these forest stands without any nutrient return. HS has an especially low mineral weathering rate because of the nutrient and base poor parent material. The input of M b cation depends almost solely on deposition rates. Table 4. Columns a-f and h contain the net proton production (>0)/proton consumption (<0) accumulated over for 24 years (1989 to 2012) [keq ha −1 24a −1 ] which is shown separately in Table 3. Column g contains remaining acid load when the M b budget (h) is not included. Columns i-n give a detailed look at the processes involved in the M b budget. M b accumulation or release (n): k + l − j − m; Acid/base reactions through M b budget (h): j + n − k. For the liming treatments, the acid load through biomass increment increases compared to the control plot because of higher incorporation of Ca and Mg in all biomass compartments and because of increase in biomass increment [22]. With regard to this calculation, it is assumed that NH4 + and NO3 − are taken up with the same proportion of 50%. In HS and AD an uptake of N only as NH4 + could almost double the acid load compared to a proportion of 50% NH4 + . In the case of IO an uptake of N only in the form of NO3 − would not only lead to a lower acid load, but result in proton consumption by biomass increment for the liming treatments ( Figure 2). The Mb cation input by mineral weathering and deposition is on all control plots not high enough to compensate the loss by seepage water and uptake by the forest stand. This indicates that whole-tree harvesting is not nutrient sustainable for these forest stands without any nutrient return. HS has an especially low mineral weathering rate because of the nutrient and base poor parent material. The input of Mb cation depends almost solely on deposition rates.

Discussion
Liming forests in Rhineland-Palatinate was performed under the assumption of rapid acidification of soils taking place by air pollutants from anthropogenic sources [43]. During the 24 years, the control plots show that the acid load noticeably exceeds the proton consumption by Mb cations. Without liming, Mb cations compensated for less than 15% of the acid load ( Table 3). As a result Ma cations, especially Al and Mn, are released, which can destabilize clay minerals [15,44], disturb plant nutrition [45][46][47], damage fine Negative values in IO stand for proton consumption by biomass increment.

Discussion
Liming forests in Rhineland-Palatinate was performed under the assumption of rapid acidification of soils taking place by air pollutants from anthropogenic sources [43]. During the 24 years, the control plots show that the acid load noticeably exceeds the proton consumption by M b cations. Without liming, M b cations compensated for less than 15% of the acid load ( Table 3). As a result M a cations, especially Al and Mn, are released, which can destabilize clay minerals [15,44], disturb plant nutrition [45][46][47], damage fine root systems [48], and reduce the activity of soil biota [49]. With regard to the three study plots, liming partially increased the acid load because of (1) a higher output of anions (SO 4 2− , NO 3 − , Org − ) with the seepage water and (2) a higher proton production through the incorporation of more cations into the biomass. However, the reduction of M a cation release below the level of the control plot shows that the higher acid load was compensated by the carbonates of the dolomitic limestone. The clay minerals were stabilized on the plots of the liming treatments whereas on the control plots an ongoing destabilization could be observed. However, this stabilization effect started to diminish for the low-dose treatments after approximately 20 years and was limited to the upper 5 cm of the mineral soil [50][51][52].
On the three sites, the loss of anions is linked with a loss of metal cations, which is equivalent to quantitative acidification of the ecosystem [14]. Negative consequences are the depletion of the soil nutrient reserves if M b cations are leached together with anions, or the pollution of spring water and other adjacent fresh water if the anions are relocated together with M a cations like Al 3+ . The high loss of SO 4 2− and aluminum in AD caused by the liming treatments indicates dissolution of aluminum sulfates due to changed pH value (cf. [53]). To a lower extent, this was also observed in IO at a soil depth of 10 cm [22]. In IO the SO 4 2− output of the control plot is already on a high level whereby liming induced none or only a slight increase. In HS, the missing effect of the liming treatments on the SO 4 2− output is caused by the soil type. Sandy soils have a lower storage capacity for SO 4 2− than more cohesive soils with a higher percentage of clay [54]. Because of the lower S pools in HS [22], there is a smaller potential of a high long-term SO 4 2− release, as observed in AD and IO.
The liming treatment 7 with the additional sulfur bound K and Mg fertilizers (Table 2) has the highest acid load due to the high SO 4 2− output. The loss of SO 4 2− led to a coupled output of Mg and Al. On the other hand, K was retained in the ecosystem. Higher K fluxes compared to the control plots could only be observed in soil depth of 10 cm, but not below the rooting zone in 60 cm soil depth. Also, the forest stands show a tendency for higher K concentrations in older needles and in needle litter compared to the other high-dose liming treatment 5 and 8 [22].
The N budget in AD is the main source for the proton production on the control plot ( Table 4). The loss of more than 5 kg NO 3 − ha −1 a −1 [22] indicates N saturation on a low level (cf. [55]). Liming increases the leaching of NO 3 − and therefore the acid load. This is an example of the risks which are associated with liming of N saturated forest areas. However, on the control plot, NO 3 − is for the most part accompanied by Al which could be a risk for biota living in adjacent freshwater ecosystems and can cause higher costs and increased technical effort to remove aluminum from drinking water [56]. After liming the absolute portion and relative proportion of Al is reduced and instead of Al more M b cations (especially Mg) are transported together with NO 3 − to the outside of the ecosystem. To improve the accuracy of the acid load by biomass increment it is important to know in which form N is taken up by the forest stand. The sample calculations for the three study areas show that uptake of N only in the form of NO 3 − reduces the acid load considerably, in the case of IO even below zero, which means net proton consumption. On the other hand, a high proportion of NH 4 + could almost double the acid load. Experiments indicate for Picea abies that a large proportion of its N uptake from the soil is NH 4 + [57][58][59]. Based on the literature, it is more likely that there is an underestimate in the acid load by biomass increment when the influence of the utilized N form is not taken into account.
For the long-term, it is unknown if liming leads to a loss of a greater amount of N if a reduction of N deposition occurs. It could be hypothesized, that the amount of N lost with seepage water is the same. Under current deposition, limed ecosystems could show a more rapid increase of N in seepage waters. However, with reduced N deposition the limed areas may begin to store N and evade N saturation at an earlier point in time. Liming increases tree growth and additional N can be retained by the vegetation or removed by harvesting.
In IO and HS, liming leads also to a slight increase of the acid load from the N budget but is less important to the other proton sources. Most of the N input is retained in the ecosystem of these two study areas even on the high-dose treatment 8 with 15 t/ha. In the case of IO liming could prevent the ecosystem to reach N saturation as there was significantly more N taken up and stored in the stem wood of the trees on the liming trials [22]. Possible reasons for this could be a higher number of living cells (cf. [60]) or that N is taken up to a greater extent as NO 3 − instead of NH 4 + , which is stored in the wood (cf. [61]). Why this effect could not be observed for the spruce forest stand in AD is not clear.
This study clearly shows the importance of minimizing the input of N into forest ecosystems as long as most of the N is still retained and not lost as nitrate by seepage water flux. Otherwise, nitrate loss would likely continue over decades similar to the sulfate outputs (cf. [3,62]).
The lower amount of anions in the seepage water in HS compared to AD and IO leads to stronger retention of the applied M b cations (Mg and Ca). Therefore, sandy soils should not be excluded automatically from liming treatments. Instead, the period between liming treatments can be longer for sandy soils with similar conditions to HS. The lower cation exchange capacity in HS [22] does not cause higher outputs of Mg and Ca by the seepage water.

Data Availability Statement:
The data presented in this study are available on request from the corresponding author.